literature review of the influence of large impoundments ... · project steering group: ian milne,...

Literature review of the influence of large impoundments on downstream temperature, water quality and ecology, with reference

to the Water Framework Directive

Scottish Environmental Protection Agency

APEM Ref: 413570

March 2015

Hannah Austin, David Bradley, Iain Stewart-Russon and Nigel Milner

Client: Scottish Environmental Protection Agency

Address: SEPA Corporate Office

Erskine Court

The Castle Business Park

Stirling

FK9 4TR

Project reference: 413570

Date of issue: March 2015

________________________

Project Director: Dr David Bradley

Project Manager: Hannah Austin

Project Steering Group: Ian Milne, Heather Ferguson, Michael Wann (Scottish Environment Protection Agency)

Graham Rutt, Katie Fawcett (Natural ResourcesWales)

Colin Gibney (Northern Ireland Environment Agency)

Judy England (Environment Agency)

________________________ APEM Ltd

Riverview

A17 Embankment Business Park Heaton Mersey

Stockport SK4 3GN

Tel: 0161 442 8938 Fax: 0161432 6083

Registered in England No. 2530851

Report should be cited as: APEM (2015) Literature review of the influence of large impoundments on downstream temperature, water quality and ecology, with reference to the Water Framework Directive. APEM Scientific Report 413570. Scottish Environment Protection Agency. pp 84.

Revision and Amendment Register

Version Number

Date Section(s) Page(s) Summary of Changes Approved by

1.0 20/02/15 Draft for comments DB

2.0 17/03/15 Draft comments addressed DB

Contents

1. Introduction ...................................................................................................................... 1

1.1 Background............................................................................................................... 1

1.2 Aims ......................................................................................................................... 2

1.3 Objectives ................................................................................................................. 2

1.4 This report ................................................................................................................ 2

2. Summary of literature search ........................................................................................... 4

3. Findings of the literature review ....................................................................................... 7

3.1 Overview ................................................................................................................... 7

3.2 Thermal and chemical characteristics of unregulated rivers and the general effects of

impoundments and natural lakes ......................................................................................... 7

3.2.1 Effects of natural lakes compared with artificial impoundments .......................... 7

3.2.2 Thermal regime – unregulated rivers ................................................................. 9

3.3 How do impoundments affect the thermal regime and water chemistry of

downstream watercourses? ............................................................................................... 11

3.3.1 Influences of impoundments on thermal regime ............................................... 11

3.3.1.1 Downstream propagation of thermal effects and recovery ............................ 13

3.3.2 Influences of impoundments on water chemistry .............................................. 19

3.3.2.1 Downstream propagation of water chemistry effects and recovery ............... 22

3.4 Influences of impoundment thermal and water chemistry effects on ecology .......... 28

3.4.1 Macrophytes and phytobenthos ....................................................................... 29

3.4.2 Macro-invertebrates ......................................................................................... 31

3.4.3 Fishes .............................................................................................................. 35

3.5 What processes affect the influence of impoundments on downstream watercourses

and their ecology? ............................................................................................................. 41

3.5.1 Physical and operational impoundment characteristics .................................... 41

3.5.2 Downstream watercourse characteristics ......................................................... 51

4. Risk assessment decision framework ............................................................................ 54

4.1 Mitigation Measures ................................................................................................ 57

5. Evaluation of Water Framework Directive environmental standards ............................... 59

6. Knowledge gaps ............................................................................................................ 60

7. Conclusions ................................................................................................................... 61

8. Recommendations ......................................................................................................... 63

9. References .................................................................................................................... 64

List of Figures

Figure 2.1 Proportion of published and grey literature identified in the literature search ...... 6

Figure 2.2 Proportion of literature from different continents identified in the literature search (other denotes documents covering multiple areas) ................................................................ 6

Figure 3.1 Factors influencing the thermal regime of rivers: reproduced from Caissie (2006) .......................................................................................................................... 9

Figure 3.2 Mean daily and diel variability of water temperatures as a function of stream order / downstream direction: reproduced from Caissie (2006) ............................................. 11

Figure 3.3 Thermal tolerance ranges (min-max, oC) for reproduction in some common British fish, showing the transition from cold water species - brown trout (Salmo trutta), Atlantic salmon (Salmo salar) and grayling (Thymallus thymallus); through cool water species – perch (Perca fluviatilis), pike (Esox lucius), roach (Rutilus rutilus) and minnow (Phoxinus phoxinus); to warm water species – bleak (Alburnus alburnus), common bream (Abramis brama), chub (Leuciscus cephalus), silver bream (Blicca bjoerkna) and tench (Tinca tinca). Adapted from Webb and Walsh (2004) ................................................................................. 36

Figure 3.4 Relative influence of stream characteristics on temperature in small, medium and large streams from Poole and Berman (2001) ................................................................ 53

Figure 4.1 Risk assessment decision framework .............................................................. 56

List of Tables

Table 2.1 List of search terms............................................................................................ 4

Table 2.2 List of search results .......................................................................................... 4

Table 3.1 Types of thermal modification downstream of impoundments .......................... 17

Table 3.2 Types of chemical modification downstream of impoundments ........................ 24

Table 3.3 Biotic responses to thermal and chemical effects downstream of impoundments – macrophytes and phytobenthos ......................................................................................... 29

Table 3.4 Biotic responses to thermal and chemical effects downstream of impoundments – macro-invertebrates ........................................................................................................... 33

Table 3.5 Biotic responses to thermal and chemical effects downstream of impoundments – fishes ........................................................................................................................ 39

Table 3.6 Impoundment characteristic risk factors for downstream water quality impacts (including occurrence of stratification) ................................................................................... 45

Table 3.7 Watercourse characteristic risk factors for downstream water quality impacts . 51

APEM Scientific Report 413570

March 2015 Page 1

1. Introduction

1.1 Background

The implementation of the European Union Water Framework Directive (2000/60/EC) (WFD) in the UK has driven the need to review the operation of many artificial impoundments of rivers and lakes, in terms of how controlled releases of water meet environmental objectives in the river water bodies downstream.

The WFD requires all water bodies to achieve good ecological status (GES), except those that have been designated as heavily modified water bodies (HMWB). HMWBs are required to achieve good ecological potential (GEP) after their economic and social values, together with the technical and financial costs of altering them, have been taken into consideration.

A water body may be designated as a HMWB where its physical characteristics have been substantially changed as a consequence of one or more of the following specific uses, and where the changes to its hydromorphological characteristics required to achieve GES would have a significant adverse impact on these uses or the water environment:

navigation, including port facilities;

recreation;

activities associated to water storage, such as hydropower generation or drinking water supply;

water regulation;

flood protection;

land drainage; or

other important sustainable human development activities,

There are lists of appropriate mitigation measures for each of the above uses. For a HMWB to achieve GEP, all relevant mitigation measures must be in place (within the technical, economic and social constraints of the impoundment operation). Mitigation measures associated with the water storage and supply uses include making controlled releases of water from impoundments.

Much attention has been devoted to understanding when and how much water should be released from impoundments to meet environmental objectives. By contrast, relatively little consideration has been given to the temperature and chemistry of water released from impoundments.

The Scottish Environment Protection Agency (SEPA) and other competent authorities in the UK have devoted a lot of attention to defining environmentally acceptable flows downstream of major impoundments. The GEP guidance provided by the UK Technical Advisory Group for the WFD (UKTAG) was updated in 2013 to reflect a better understanding of environmental flow requirements (UKTAG, 2013). In order to be able make the right decisions about mitigating unacceptable environmental impacts in rivers downstream of impoundments, competent authorities need to understand the potential ecological consequences arising from not only the quantity of water released, but also from the quality of the water released. More information is needed on how large impoundments affect the quality of water released in order to avoid any unintended negative consequences to the ecological status of downstream water bodies resulting from the introduction of mitigation measures involving water releases.


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The need for this project arose from SEPA investigating potential mitigation measures involving deep-water releases from a large impoundment to achieve WFD objectives for water quantity based on current guidance provided by the UK Technical Advisory Group (UKTAG, 2013; UKTAG, 2008a).There was no specific guidance available regarding how such environmental flows are best delivered (e.g. surface water vs deep water release), nor the potential trade-offs between delivering such flows and their potential water quality (temperature and chemistry) implications for downstream water bodies.

1.2 Aims

In order to assist the UK regulatory agencies manage flow releases downstream of impoundments to ensure compliance with environmental standards, this project aims to review existing information on the effects of large impoundments on downstream water quality (including temperature and chemistry) and ecology, with an emphasis on information relevant to the UK. It aims to consider the implications for the WFD; specifically, changes to ecological potential/status in water bodies that might be ‘re-wetted’ with the introduction of deep water releases from large impoundments.

1.3 Objectives

The specific objectives of this report are:

to undertake a literature review on the nature and scale of effects of large impoundments on downstream river temperature and other aspects of water quality, and the resulting effects on river ecology;

assess and document the conditions leading to significant impacts, and factors which influence impact, such as size of impoundment and local climate;

consider the extent to which it may be possible to predict, from impoundment characteristics (such as, volume, depth and location), the likely scale of ecological impact from physico-chemical issues under a deep-water re-wetting scenario;

consider whether the evidence from the literature review supports use of the existing WFD environmental standards for water temperature and dissolved oxygen in relation to HMWBs and the relevant mitigation measures required to achieve good ecological potential, and;

identify any knowledge gaps, with particular reference to the UK.

1.4 This report

This report is structured as follows:

Section 2 presents a summary of the literature search including methods and outcomes;

Section 3 presents the findings of the literature review: o Section 3.1 provides an overview of the literature review; o Section 3.2 introduces general background to the effects of natural lakes and

impoundments on downstream watercourses, as well as the thermal regime of unregulated rivers;

o Section 3.3 describes the mechanisms driving water quality effects downstream of impoundments and their downstream persistence;

o Section 3.4 describes potential knock-on effects on ecology;


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o Section 3.5 explores the characteristics of impoundments that may be indicators of significant downstream effects;

Section 4 presents a risk-based decision framework to assist with prioritising impoundments for further investigation and selecting appropriate mitigation measures;

Section 5 includes an evaluation of relevant WFD environmental standards in light of the literature review findings;

Section 6 presents knowledge gaps;

Conclusions and recommendations are presented in Sections 7 and 8; and

References are presented in Section 9.


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2. Summary of literature search

A global scientific literature search was undertaken in October and November 2014 using computerised searches of the ISI Web of Knowledge which includes the following databases: Web of Science (1990-), BIOSYS Citation Index (1969-), BIOSYS Previews (1969-), Data Citation Index (1900-), MEDLINE (1950-) and Journal Citation Reports. The search was based on combinations of the search terms that were discussed and agreed with the project steering group (Table 2.1). The search terms were combined to include the dam and/or its release, plus the effect on downstream water temperature, water chemistry and/or ecology.

Table 2.1 List of search terms

A) Relating to the dam and its

releases

B) Relating

to the ecology

C) Relating to downstream water

chemistry/temperature

D) Relating to reservoir

water quality

E) Specific water

chemistry terms

F) Water Framework Directive / Policy

Impoundment Ecology Water quality Stratification Iron Mitigation measures

Dam Fish Water temperature Hypolimnion Manganese Heavily Modified Waterbody

Reservoir Thermal regime Hypolimnetic Monitoring Guidance

Reservoir release Temperature variability Epilimnion River Continuity

Compensation flow Thermal variability Epilimnetic

Environmental flow Dissolved oxygen

Freshet

Regulated river

Re-wetting

The search term combinations and the number of results returned (total 16,767) are presented in Table 2.2. All searches were undertaken by a single reviewer, and a total of 2062 results were assessed for suitability (searches with number of results greater than 300 were excluded), based on whether they focused on abiotic (temperature and water chemistry) impacts of impoundments on downstream watercourses, with a focus on temperate climate areas in general, and the UK in particular.

Table 2.2 List of search results

Search Term No. of results

"compensation flow" reservoir 11

compensation flow reservoir 175

reservoir release fish temperature 164

environmental flow freshet 65

dam mitigation ecology 154

reservoir fish temperature 1538

compensation flow fish temperature 20

impoundment fish temperature 243

dam fish temperature 917

environmental flow fish temperature 2265

flow regime fish temperature 442

freshet fish temperature 31

regulated river fish temperature 283

re-wetting fish temperature 0

rewetting fish temperature 2


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Search Term No. of results

hydropeaking fish temperature 18

Reservoir release ecology 1272

Reservoir release water quality 626

Reservoir release ecology water quality 192

Reservoir release ecology temperature 249

Reservoir release ecology dissolved oxygen 61

Flow regime ecology 5870

Flow regime water quality 1329

Flow regime dissolved oxygen 282

Flow regime ecology water quality 446

Release flow regime ecology water quality 38

Reservoir release ecology water quality epilimnion epilimnetic 0

Reservoir release ecology water quality hypolimnion hypolimnetic 5

reservoir release thermal regime environmental flow ecology 15

reservoir release thermal variability environmental flow ecology 5

reservoir release temperature variability environmental flow ecology 5

reservoir release thermal regime environmental flow ecology river continuity 0

reservoir stratification hypolimnetic hypolimnion manganese iron 4

heavily modified waterbody mitigation measures reservoir release temperature water quality 0

heavily modified waterbody mitigation measures reservoir release temperature water quality monitoring guidance 0

heavily modified waterbody 3

heavily modified waterbody mitigation measures 0

heavily modified water body mitigation measures 3

heavily modified water body reservoir release 1

Reservoir release ecology water quality epilimnion 3

Reservoir release ecology water quality epilimnetic 4

Reservoir release ecology water quality hypolimnion 15

Reservoir release ecology water quality hypolimnetic 11

Gillespie et al. (2014) note that their recent literature search revealed inconsistent use of terms and keywords used to describe research concerning the impact of reservoir flow modification on downstream ecological conditions, a finding that is borne out by this study. To aid efficiency of future literature searches, Gillespie et al. (2014) recommended that all future literature concerning this topic includes the keywords “environmental flow” and “reservoir” where possible. If adopted, such a change is considered likely to improve the efficiency of future searches of this nature.

Once the key search terms had been used, the agreed terms were modified to provide a wider search for relevant literature. A grey literature search was undertaken in December 2014 using Google Scholar, and was also based on appropriate combinations of the search terms presented in (Table 2.1). Some papers and grey literature reports were provided directly by the project steering group. The initial list of search results (2) was screened by two technical reviewers using their expert opinion to identify only literature sources specifically relevant to this study. Search results were screened based on their abstracts where available, and based on their titles where abstracts were not available. Following this screening process, the list of references was refined to 185 relevant primary sources, comprising 165 published references and 20 grey literature sources (Figure 2.1). The primary sources identified from both peer-reviewed and grey literature sources are presented in Section 9, together with further references identified from several important review papers. The relevant references were collated and reviewed.


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Figure 2.1 Proportion of published and grey literature identified in the literature search

The search identified most primary sources from North America, the UK and the rest of Europe, and relevant literature was also identified from other continents including North and South America, Asia, Australia and Africa (Figure 2.2). General patterns and the physical processes behind the effects of large impoundments on downstream water temperature, water chemistry and ecology were investigated, based on global literature with a focus on issues of interest to the UK.

Figure 2.2 Proportion of literature from different continents identified in the literature search (other denotes documents covering multiple areas)


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3. Findings of the literature review

3.1 Overview

From the global literature it is clear that impoundments throughout the world interrupt the river continuum and can exert a profound influence on downstream water chemistry, thermal regime and ecology (Petts, 1986; Caissie, 2006). Although such effects are well recognised (e.g. Ward and Stanford, 1982; Ward, 1985; Petts, 1986; Webb and Walling, 1996), they are examined much less often than river hydrology (Olden and Naiman, 2010).

There has been increasing focus on environmental flow releases from impoundments, with general agreement that some resemblance to natural flow variability is beneficial to support a naturally functioning ecosystem. However, although there is ongoing investigation and debate regarding how best to define the quantitative aspects of environmental flows, to date there has been less explicit consideration of the water quality components including temperature – a fundamental ecological variable (Olden and Naiman, 2010). Every reservoir modifies the physicochemical parameters of water – the degree of modification depends on the size and function of the reservoir and the environment in which it is located (Brooker, 1981; Soja and Wiejaczka, 2014).

Based on the sources identified from the literature search, the following sections describe some general background to the effects of natural lakes and impoundments on downstream watercourses, and the thermal regime of unregulated rivers; the mechanisms driving water quality effects downstream of impoundments and their downstream persistence; potential knock-on effects on ecology; and the characteristics of impoundments that may be indicators of significant downstream effects.

3.2 Thermal and chemical characteristics of unregulated rivers and the general effects of impoundments and natural lakes

Influences on the thermal regime in unregulated rivers are described below, together with the effects of natural lakes, to put the effects of impoundments into context.

3.2.1 Effects of natural lakes compared with artificial impoundments

The limnological processes in a reservoir are often relatively similar to those in a natural lake; Brooker (1981) reports that the chemical and thermal changes resulting from impoundment are generally similar to those reported for natural lakes (Hutchinson, 1975; Wetzel, 1975) and that even short term storage of water can result in substantial changes in chemistry (Slack, 1978). In broad terms, both lakes and reservoirs constitute a large body of water with a high thermal capacity, lower surface area to volume ratio and a longer retention time compared to a natural stream (Petts, 1986). The relatively high thermal capacity and low surface area to volume ratio of lakes and reservoirs means that water temperature in the lentic water body is less variable in response to daily fluctuations in solar radiation and annual climatic variations.

In lentic or slow-flowing water bodies during the spring and summer, the upper layers of the water may warm due to increasing ambient air temperatures. Above 4oC water becomes less dense as it warms and so a warmer water layer (epilimnion) may develop above a layer of cooler, denser water (hypolimnion).Thermal stratification is the development of a pronounced barrier (thermocline) between the warmer and cooler layers of water (due to the


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relationship between water density and temperature) in the absence of sufficient mixing. Thermal stratification may occur over very short or longer timescales depending on the occurrence of mixing (e.g. due to high wind speeds). Where stratification is continuously present for several months of the year it may be termed ‘stable’. Typically as surface waters cool during autumn they become denser and are replaced by warmer waters from below; this characteristic autumn overturn creates fully mixed conditions throughout the water column (Petts, 1986).

The development of stratification in natural lakes and reservoirs is dependent on the lake/reservoir characteristics. Typically deep water bodies with a low surface area to volume ratio, where the influence of inflows/outflows and wind-induced mixing is low, are more prone to stratification (Brooker, 1981). In general, the shorter the water body retention time, the greater the influence of inflows on thermal behaviour and the more exposed the water body, the greater the degree of wind-induced mixing (Petts, 1986). This will increase water column mixing and help break down or limit the development of stratification.

Eutrophication may also occur in both impoundments and natural lakes, influenced by catchment land-use and other inputs (e.g. from lake bed or reservoir sediment). This may affect the chemistry of water leaving either the natural lake via its outflow, or the impoundment via its discharge.

However, there are several key differences between natural lakes and some types of impoundments which mean that their respective effects on downstream watercourses can differ:

The conditions leading to stable stratification may be more common in impoundments than in natural lakes since reservoirs tend to be constructed so as to minimise the land area drowned by the impoundment whilst maximising the impounded volume of water, and therefore they are often deep and steep-sided, with a low surface area to volume ratio which can inhibit the influence of wind-induced mixing (Petts, 1986).

Whilst the outflows from natural lakes almost invariably occur from the lake surface (except where the outflow runs underground e.g. in limestone areas), discharges from impoundments can be made from any depth (depending on the construction of the impoundment, its purpose and mode of operation). Impoundments frequently release water from depth rather than from the surface. This can affect water movement within the reservoir, as well as the thermal and chemical characteristics of water released.

Because reservoir water levels are generally more variable than those of a natural lake, advective heat transfer and the associated vertical mixing tends to be more important than within a natural lake (Petts, 1986).

Impoundments may influence downstream water quality (temperature and chemistry) both directly and indirectly (Preece, 2004). Direct changes occur when the temperature or chemistry of released water differs markedly from what would naturally be expected in the downstream watercourse. Direct changes may occur in both lakes and impoundments, although certain types of direct changes (e.g. those associated with stratification) may be more prevalent in impoundments due to their dimensions or operational regime.

Impoundments can also affect downstream temperature and water chemistry indirectly by influencing processes that control the delivery, distribution and retention of heat and chemicals within the downstream river channel (Preece, 2004). For


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example, the quantity and dynamics of flow released from an impoundment will affect its capacity for heating or cooling, and may alter the proportion of water originating from thermally contrasting sources in comparison with the natural scenario (Preece, 2004; Petts, 1986; Poole and Berman, 2001). Such indirect effects are distinct to impoundments in contrast to natural lakes (which by definition have a natural discharge regime).

Although not the main focus of this study, it is also important to note that severe thermal fluctuations over short periods can also be caused by hydro-peaking releases resulting from power generation (Brooker, 1981). Pfitzer (1967) reported summer temperatures rapidly changing by 6-8oC below deep release dams as power releases were required. Changes in outlet valves may also produce rapid thermal changes below reservoirs.

For the purposes of this study, the effects of natural lakes on downstream water bodies are only of interest inasmuch as they highlight the effects of man-made impoundments. Therefore, although stratification may develop in natural lakes, the discussion below focuses on impoundments.

3.2.2 Thermal regime – unregulated rivers

For streams unaffected by upstream impoundment or regulation, there are a multitude of factors which influence the thermal regime which can be classified into four broad categories (Caissie, 2006) (Figure 3.1):

atmospheric conditions;

topography;

stream discharge; and

streambed.

Figure 3.1 Factors influencing the thermal regime of rivers: reproduced from Caissie (2006)


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Caissie (2006) notes that: atmospheric conditions (in particular solar radiation) are among the most important factors and are mainly responsible for the heat exchange processes that take place at the water surface; topography / geographical setting is also important because it influences atmospheric conditions; and stream discharge (mostly a function of river hydraulics) mainly influences the heating capacity and/or cooling through mixing of water from different sources (including streambed heat exchanges). It is important to note that the thermal capacity of water is very high in comparison with other substances (e.g. air), meaning that it can absorb a large amount of heat before its temperature increases substantially (Gordon et al. 2004).

On the spatial scale, water temperature tends to increase or decrease in the downstream direction to a point where it reaches equilibrium with air temperatures and other influencing factors such as groundwater and tributary inflows. According to Bloomfield et al. (2013), the temperature of shallow groundwater is reported as generally 1-2°C higher than the mean annual surface temperature (Busby et al. 2009), where mean annual air temperature at sea level in the UK varies from 8°C in the north to 12°C in the south. In northern temperate climate regions the temperature of shallow groundwater varies seasonally and diurnal variations may be seen above 1.5 m depth (Busby et al. 2009; Taylor and Stefan, 2009).

The relationship between river water temperature and its influencing factors is non-linear: the rate of downstream increase tends to be greater for small streams (and particularly those that are wide and shallow) than for large and deep rivers; a smaller, shallower body of water is more easily affected by atmospheric conditions than a large one (Caissie, 2006) (Figure 3.2). Maximum temperatures typically exhibit a progressive increase from the headwaters to the mouths of river systems and annual ranges also tend to increase with distance downstream (although the lower reaches of large rivers may sometimes exhibit slightly smaller annual ranges than segments immediately upstream (Caissie, 2006)).

Both the annual range and annual mean temperature tend to decrease with increasing altitude, which exerts a primary influence on the thermal regime of running waters through its influence on air temperature. However, altitude is only one of several factors that may be responsible for downstream changes in temperature conditions; aspect and canopy (both related to insolation) were found to mask the effect of altitude at two of the sites studied by Johnson (1971).

On the temporal scale, river water temperature naturally varies on both diel and annual cycles, also showing considerable inter-annual variability. On a daily timescale, riverine water temperature minima generally occur in the early morning (NB this can vary throughout the year and may occur later during the winter months and further downstream within the catchment (Ward, 1985)), with maxima occurring during the late afternoon and early evening. The daily variability in water temperature is generally smallest for groundwater-dominated headwater streams and large, deep rivers with a high thermal mass. The daily water temperature variability tends to be highest for wide and shallow rivers in which the water is more readily influenced by prevailing meteorological conditions (solar radiation, air temperature etc.) (Caissie, 2006) (Figure 3.2).

On an annual timescale, riverine water temperature in temperate regions (such as the UK) follows a sinusoidal function with minima generally occurring in natural streams during late winter (typically February) and maxima in mid-summer (typically July) (Webb and Walling, 1996). The proportion of groundwater baseflow vs. surface water flow, as well as position in the catchment and prevailing atmospheric conditions will all affect the timing of annual


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minima and maxima, and considerable inter-annual variability has been observed (Webb et al. 2008). Considerable inter-year variation is also apparent within particular catchments (Webb and Walling, 1996).

Figure 3.2 Mean daily and diel variability of water temperatures as a function of stream order / downstream direction: reproduced from Caissie (2006)

3.3 How do impoundments affect the thermal regime and water chemistry of downstream watercourses?

General effects of impoundments on the downstream thermal regime and water chemistry are described below, with particular reference to the WFD physico-chemical elements. Greater focus is placed on the thermal regime, and on the baseflow component of the flow regime, since these are primary project objectives.

3.3.1 Influences of impoundments on thermal regime

Ward and Stanford (1979), cited by Brooker (1981) recognised six categories of thermal modification downstream of impoundments. These are described in Table 3.1, together with a list of literature sources. Focussing on the most severe effects of continuous releases:

In deep, stratified reservoirs with a hypolimnial release, the reservoir will typically discharge relatively cold water during the summer, potentially depressing the annual


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maximum temperature, delaying the occurrence of the annual maximum, and reducing both diurnal and annual variation (Soja and Wiejaczka, 2014). In contrast, winter releases may be warmer on average than for the un-impounded river, changing the timing and occurrence of ice formation downstream (Webb and Walling, 1993).

In small and/or shallow reservoirs, or in stratified reservoirs with a surface water (epilimnial) release, the reservoir will typically discharge warm, well oxygenated water with few nutrients (potentially with high algal content) (Petts, 1986), causing elevated mean temperatures during the spring and summer months (Olden and Naiman, 2010, Lessard and Hayes, 2003) and potentially hastening the occurrence of ecologically relevant temperature thresholds in comparison with the un-impounded condition. Both diurnal and annual variation may also be reduced.

Unexpected changes in micro-climatic conditions may also occur as an indirect consequence of impoundment: Caissie (2006) highlights a study by Troxler and Thackston (1977) where the cooled air resulting from water release in a valley promoted the formation of fog, which reduced natural heat exchange between the river and the atmosphere.

In an example of the effects of a deep reservoir release, Lavis and Smith (1972) reported that maximum stream temperatures in the River Lune below Selset and Grassholme reservoirs in Yorkshire were depressed by as much as 12oC in summer. This situation is slightly complicated by the fact that the river is influenced by two reservoirs acting in series, but it is interesting to note that the draw-off from Grasshome is reported as being from depth; 28.3 m below top water level (with a depth at top water level of 29.6 m). At Lyn Brianne (another deep reservoir (85 m) with a deep water release (65 m below top water level)), the discharge regime reportedly causes a reduction in the amplitude of weekly and monthly fluctuations, a delay and reduction in the summer peak and a delay and increase in the winter temperature immediately downstream of the dam compared to further downstream (Wyke, 1997).

Webb (1995) reports that the main effect of Wimbleball Lake (a reservoir less than 50 m deep with multi-level draw-offs within the top 25 m) on temperature in the downstream River Haddeo has been to increase the mean value, eliminate freezing conditions, depress summer maxima and reduce diurnal fluctuation (albeit with considerable inter-year variability).

It is interesting to note that, investigative monitoring of the Altnahinch dam in Northern Ireland (Murphy, 2011) indicated temperature effects in the downstream watercourse even in the absence of stratification (average downstream river temperatures for July and August were 2oC higher than those measured upstream). The average depth of the reservoir is only 7-8 m and the apparent lack of stable stratification was attributed to the large surface area of the lake in comparison with its volume and the fact that it is relatively exposed with little tree cover (altitude 240 m AOD). Despite the fact that the compensation flow is drawn from 18 m depth immediately behind the dam, the fact that the reservoir is relatively shallow and fully mixed appears to cause temperature changes similar to those described in the second bullet point above.

Higher downstream temperature minima (in comparison with an upstream control site) were attributed either to spill events consisting of warmed reservoir surface water (on such occasions the lowest downstream water temperatures were generally 2oC and on one occasion 4oC higher than those measured upstream, although the highest temperatures


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remained similar), or to the effect of the heat retention characteristics of the concrete dam structures as the compensation flow passed over them (rather than to the effects of a hypolimnial release during winter). Reduced magnitude of diurnal temperature variation was observed downstream (in comparison with upstream of the dam).

It should, also be noted that as well as spatial variation in downstream thermal regime effects, temporal variability is also highly likely to occur (e.g. on daily, seasonal and inter-annual timescales). For example, inter-year variation is evident in the development of stratification at Cow Green; in 1975 under conditions of high air temperature and low wind speed, stratification did develop (the reservoir was stratified during the whole of July and August 1975) but was very rare at other times in the historical record (Crisp, 1977). Webb and Walling (1996) caution that because many thermal regime studies are based on relatively short term datasets, they may misinterpret how the downstream thermal impact of a reservoir varies in response to inter-annual fluctuations in hydro-meteorological conditions or changes in reservoir operation. The effects of a newly developed reservoir are also likely to differ from those of a reservoir that has been in place for many years (Petts, 1986).

3.3.1.1 Downstream propagation of thermal effects and recovery

Thermal impacts can encompass relatively short or long distances below their respective dams depending on heat exchange with the atmosphere, hydrologic inputs from tributaries and groundwater recharge, and dam discharge (Palmer and O’Keeffe, 1989, cited by Olden and Naiman, 2010). Webb (2008) notes that the persistence downstream of thermal modification below an impoundment will depend on the volume and temperature of water released from the reservoir, the temperature and discharge of tributary inflows and the magnitude of heat exchange between the river, ground and atmosphere (Webb, 1995). Results from the latter study (Webb, 1995), suggest that recovery distance depends upon which facet of the thermal regime is being considered, varies with the volume of flow from the reservoir and is more strongly influenced by heat exchange with the atmosphere than by tributary inflow. Webb (2008) also notes that investigations of the downstream persistence of temperature modifications has highlighted complicated patterns of recovery in systems regulated by multiple dams of different types (O’Keeffe et al. 1990).

Olden and Naiman (2010) note that scientists have provided much information concerning the extent to which the thermal effects persist downstream beyond the immediate vicinity of dams, for example:

In the Murray-Darling River Basin (Australia), depressed summer temperatures extend up to several hundred kilometres downstream of dams on a number of major rivers, including Murrumbidgee, Marquarie, Mitta Mitta, Namoi and Murray (Ryan et al. 2001; Preece, 2004; Todd et al. 2005).

The thermal recovery of the River Svratka, Czech Republic, required over 40 km (Ward, 1985), and using rates of maximum summer warming, Stevens, Shannon and Blinn (1997) estimated that a main-stem distance of 930 km would be required for water temperatures to fully recover (increase to pre-dam conditions) below Glen Canyon Dam in the Colorado River, U.S.A. (a distance that is currently prevented by other downstream dams).

In cool water systems affected by epilimnetic-release dams, streams may not be able to shed added heat during the summer and downstream water temperatures may continue to warm due to normal stream processes. Cited by Lessard and Hayes (2003), Fraley (1979) found significant summer temperature increases in the


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Madison River, Montana (USA) that never returned to upstream temperatures even 56 km downstream of a surface-release dam, although diurnal temperature fluctuations did recover with increasing distance.

Lowney (2000) reports that steady hypolimnial releases with minimal diurnal variation in water temperature can cause a characteristic pattern of ‘nodes’ of minimal diurnal variation and ‘antinodes’ of maximum diurnal variation to form at regular intervals downstream of the impoundment; nodes at locations equivalent to multiples of 24 hrs travel downstream of the impoundment, and antinodes at locations equivalent to odd multiples of 12 hrs travel downstream1. Nodes of minimum diurnal variation are essentially reproductions of the temporal signal at the impoundment. Changes in flow which affect travel time interrupt the formation of nodes and antinodes as do external sources of heat such as tributary inflows. Nodes and antinodes have been observed in the Klamath River and the Sacramento River, and similar behaviour has been observed in other regulated systems (Lowney, 2000). Thus for any monitoring regime designed to test such effects, consideration of time of travel is of critical importance when locating monitoring sites.

Meta-analysis by Haxton and Findlay (2008) suggested a large effect of a hypolimnetic release on downstream aquatic communities, and that the effect can extend some distance (+70 km) downstream before recovery occurs (citing Edwards, 1978 and Lessard and Hayes, 2003).

However, the majority of this literature describes larger impoundments and different climatic conditions to those found in the UK.

With reference to UK reservoirs, Brooker (1981) considered that the thermal and chemical effects of impoundments were generally restricted to those reaches of river immediately downstream of the impoundment. In hot climates water temperatures equilibrate rapidly with air temperatures and this is generally the case in temperate climates where the water passes through rapidly flowing, turbulent reaches.

In the UK context (reservoirs up to 200x106 m3 capacity), temperature effects have been generally described as being fairly localised, with downstream temperature effects probably becoming negligible within 10-30 km of the point of release (Crisp (1995) cited by Ellery and Wilkins (1999)). The released water will adjust towards air temperatures as it travels

1This pattern is qualitatively described by considering two parcels of water leaving the reservoir during

summer months, one at sunset and one at sunrise (Lowney, 2000). The first parcel having departed the reservoir release at sunset will initially be subjected to a night-time energy budget, its temperature changing in response to heat flux at the air-water interface, rising or falling depending upon its initial temperature. Roughly 12 hours later, the same parcel will then encounter a daytime energy budget, warming or cooling depending upon its temperature. Meanwhile, the second parcel of water departs the reservoir at sunrise and is first exposed to a daytime and then to a night-time energy budget. Because both parcels of water depart the reservoir at the same temperature, they arrive at a location equivalent to 1 day's travel time downstream of the reservoir having gained or lost roughly the same amount of heat, provided that meteorological conditions do not change dramatically from day to day. The first antinode is formed at 12 hours travel time from the reservoir release and illustrates the temperature difference between the response of the two parcels to daytime and night-time energy budgets (Lowney, 2000).


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downstream, and it is probable that this will be more rapid with the addition of tributaries (Ellery and Wilkins, 1999).

A number of UK examples are cited in the literature with varying degrees of downstream propagation and recovery. These are presented below in order of reported extent of downstream effect (longest to shortest distance downstream):

Haile, James and Sear (1989) reported that winter temperatures in the River North Tyne downstream of Kielder Water (vol. ~199 Mm3) were significantly warmer than ambient conditions, for example, during a winter week 1.5oC above the dam at Butteryhaugh, compared with 4.5oC immediately below the dam at Yarrow and 2.5oC at Chollerford 35 km downstream of the reservoir. In contrast, in June, the temperature below the dam was still only 7oC when at Chollerford it had reached 15oC. The authors concluded that there was ecological impact for at least 14 km downstream of the dam.

In an investigation of the Afon Clywedog (vol. ~50 Mm3) and Afon Vyrnwy (vol. ~60 Mm3), Cowx et al. (1987) note that although seasonal and diel temperature variations are markedly influenced by release discharge and depth of withdrawal, the downstream influence of river regulation is confined to a relatively short reach (~30 km) below the dams.

Wyke (1997) notes that the impact of the discharge regime from Llyn Brianne (vol. ~64 Mm3) on downstream river temperature is dampened with distance downstream of the dam, because the water re-equilibrates with air temperature and is diluted by the inflowing of unregulated tributaries. Effects were lower at a monitoring site ~20 km downstream than immediately below the dam. Wightman et al.(1990) note that under normal summer conditions, reduced river temperatures downstream of Llyn Brianne extended as far as Llandovery (~20 km downstream), although extension much further downstream may occur during larger scale releases. A river management (freshet) release in the summer of 1983 reduced river temperatures from 20oC to 13oC at Manorafon, 36 km downstream of the reservoir. A recent survey in July 2013, using a high resolution thermal imaging camera mounted in an aircraft, showed a cooling effect of Llyn Brianne up to 25 km downstream of the impoundment (APEM, 2013).

Data collected by Summers (undated a and b) indicates that downstream of Loch Lyon (vol. ~59 Mm3), temperature effects had attenuated by 16 km (10 miles) and more or less returned to naturalised by approximately 25 km downstream of the dam. Jackson et al. (2007) concur that macro-invertebrate communities in the River Lyon begin to resemble those that might be expected in the unregulated river (as referenced by the Lochay) within 10 km downstream of the lower dam at Meggernie.

Temperature effects downstream of Caban Coch (vol. ~36 Mm3) in Wales are reported to have little effect on the River Wye after ~20 km downstream of the dam (Brooker, 1981 and Ellery and Wilkins, 1999).

Webb (1995) reports that impacts of Wimbleball Lake (vol. ~22 Mm3) on temperature in the downstream River Haddeo persisted over a distance of at least 5 km below the dam but declined downstream, especially for seasonal and diurnal variations. It is also noted that regulation has affected the thermal regime of the downstream River Exe, with the influence generally restricted to a distance of up to 20 km from the dam, but during conditions of hot weather and low flows extending to almost 40 km downstream.

Crisp (1977) notes that little attention has been paid to the attenuation of thermal and chemical effects downstream of Cow Green Reservoir (vol. ~41 Mm3) on the River


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Tees because the changes will be diluted by the entry of major tributaries not far below the dam; data are presented only for the reach immediately below the dam.

Murphy (2011) concluded that Altnahinch Dam (vol. ~1.1 Mm3) impacted the river system downstream as far as the confluence with the Lewin Burn which is approximately 2.5 km.

Lessard and Hayes (2003) found that in general, mean summer temperature was substantially increased downstream of small, surface release facilities, and that the increases were maintained at least 2-3 km below the dams.

Tosney (2013) noted minimal effects of scour releases on downstream temperature in Yorkshire, although scour releases were only undertaken in spring and autumn in this study.

These data largely confirm the conclusions of Crisp (1995) that in the UK, downstream temperature effects probably become negligible within around 10-30 km of the point of release. It is also interesting to note that reservoir volume does not appear to be the main driver of downstream extent of effect.


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Table 3.1 Types of thermal modification downstream of impoundments

Process / implication Influence and mechanism Risk factors Sources

Reduced magnitude of diurnal temperature variation (increased diurnal constancy)

Impoundments generate a large, lentic body of water with a high thermal capacity compared to the natural stream. The water temperature is less easily influenced by variations in solar radiation and air temperature. Thus, water released downstream tends to exhibit less temperature variability when measured on a sub-daily timescale. This is highly susceptible to the downstream measurement location c.f. Lowney (2000) due to the development of nodes and antinodes of minimum and maximum diurnal variation.

Any natural lake or impoundment may cause this effect due to the large volume of water retained (Very Certain) Exacerbated by*: Large volume Long retention time Stable stratification (low exposure to wind, high depth, low surface area to volume ratio) Hypolimnial releases Epilimnial releases Low Base Flow Index (BFI); low volume and/or distant tributary inputs; riparian vegetation/tree cover; over-deepened / narrowed channel; other anthropogenic factors e.g. water abstraction.

Lowney (2000) Brooker (1981) citing Ward and Stanford (1979) Lavis and Smith (1972) Soja and Wiejaczka (2014) Webb (1995) Webb and Walling (1988a, 1996, 1997, 1998) Summers (undated a and b) Wyke (1997) Jackson et al. (2007) Crisp (1977) Crisp et al. (1983) Cowx et al. (1987) Murphy (2011) reported possible reduced variation from a shallow impoundment (mean depth 7 m), mechanism unknown. Lessard and Hayes (2003) reported reduced variation from “small”, surface release dams.

Reduced magnitude of seasonal temperature variation (increased seasonal constancy)

Impoundments generate a large, lentic body of water with a high thermal capacity compared to the natural stream. The water temperature is less easily influenced by meteorological / climatic conditions during the year. Thus, water released downstream tends to exhibit less temperature variability when measured on a seasonal timescale. This is especially the case where a stable stratification develops, and water is released from a thermally stable hypolimnion or epilimnion. NB even without stratification, there will be an effect of impoundment on downstream temperature variability (particularly diurnal variability) due to the higher thermal capacity of the impounded water (even the epilimnion as evidenced by Lavis and Smith (1972)).

Reduction in mean and / or maximum summer temperatures (summer depression)

The depth from which water is withdrawn is particularly important in determining the temperature of water discharged downstream. The temperature of water discharged from many large and deep impoundments in summer (particularly at depth) tends to be cooler than that in the natural stream. Where stratification occurs and water is drawn from the hypolimnion, this effect is emphasised since there is no atmospheric heating of the hypolimnion whilst stratification prevails.

Deep impoundments; hypolimnial releases (Certain) Exacerbated by: Long retention time Stable stratification (low exposure to wind, high depth (>5 m), low surface area to volume ratio) Low BFI; low volume and/or distant tributary inputs; riparian vegetation/tree cover; over-deepened / narrowed channel; other anthropogenic factors e.g. water abstraction.

Brooker (1981) citing Ward and Stanford (1979) Webb and Walling (1988a, 1996, 1997, 1998) Summers (undated a and b) Lavis and Smith (1972) Soja and Wiejaczka (2014) Jackson et al. (2007) Crisp (1977) Cowx et al. (1987) Wightman et al. (1990) Halleraker et al.(2007)

Increase in mean and /or minimum winter temperatures (winter elevation)

Where stratification occurs the depth from which water is withdrawn is particularly important in determining the temperature of water discharged downstream; in winter, water drawn from the hypolimnion tends to be warmer than that in the natural stream.

Deep impoundments; hypolimnial releases (Certain) Exacerbated by: Low BFI; low volume and/or distant tributary inputs; riparian vegetation/tree cover; over-deepened / narrowed channel; other anthropogenic factors e.g. water abstraction.

Brooker (1981) citing Ward and Stanford (1979) Webb and Walling (1988a, 1996, 1997, 1998) Haile, James and Sear (1989) Soja and Wiejaczka (2014) Jackson et al. (2007) Cowx et al. (1987) Wightman et al. (1990) Halleraker et al. (2007)

Increase in mean and /or maximum summer temperatures (summer elevation)

In small or shallow impoundments (or deeper impoundments where water is drawn from the epilimnion), the temperature of water discharged may be warmer than that in the natural stream. This effect may be emphasised where stratification occurs since all atmospheric heating occurs in the epilimnion, and there is no mixing with the cooler water of the hypolimnion whilst stratification prevails.

Small and/or shallow impoundments (Quite Certain) Epilimnial releases from deep impoundments (Quite Certain) Natural lakes Exacerbated by: Low BFI; low volume and/or distant tributary inputs; riparian vegetation/tree cover; over-deepened / narrowed channel; other anthropogenic factors e.g. water abstraction.

Brooker (1981) citing Ward and Stanford (1979) Murphy (2011) Lessard and Hayes (2003) Fraley (1979) cited by Lessard and Hayes (2003)


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Process / implication Influence and mechanism Risk factors Sources

Change in rate and/or timing of ecologically relevant temperature increases, decreases and thresholds, either delays or advancements (thermal pattern change)

Because of the higher thermal capacity of impounded water, the rate of change of water temperature in the reservoir is slower than that in the natural stream – the large volume of water takes longer to heat and cool down, respectively, in response to changing atmospheric conditions in spring and autumn. This can influence the timing of achievement of temperature thresholds in the river downstream (by either delaying or advancing the date on which they occur). Because they are a source of water with a very stable temperature profile, impoundments can also affect the rate of change of temperature in the downstream watercourse. Variable, rapid and untimely releases due to hydro-peaking can cause rapid changes in downstream thermal regime. NB if reservoir retention times are very short, these effects are likely to be lessened as the reservoir outflow will be increasingly influenced by the characteristics of the reservoir inflows. Impacts associated with rapid and untimely fluctuations will not occur where such releases do not occur.

Any natural lake or impoundment may cause this effect due to the large volume of water retained, although such effects are considered “natural” in the context of the presence of natural lakes. Exacerbated by: Hypolimnial releases (Quite Certain) Hydro-peaking releases causing rapid and untimely fluctuations in temperature (Uncertain) Long retention time Stable stratification (low exposure to wind, high depth (>5 m), low surface area to volume ratio) Epilimnial releases Low BFI; low volume and/or distant tributary inputs; riparian vegetation/tree cover; over-deepened / narrowed channel; other anthropogenic factors e.g. water abstraction.

Steady releases: Brooker (1981) citing Ward and Stanford (1979) Webb and Walling (1988a, 1996, 1997, 1998) Haile, James and Sear (1989) Soja and Wiejaczka (2014) Wyke (1997) Crisp (1977) Cowx et al. (1987) Wightman et al. (1990) Hydropeaking releases: Stanford and Hauer (1992) note temperature fluctuations associated with regulation releases. Ellery and Wilkins (1999) note a greater effect of regulation releases vs compensation) Tosney (2013) noted minimal effect of scour releases on downstream temperature in Yorkshire, although scour releases were only undertaken in spring and autumn in this study.

* NB the inverse of exacerbating factors may act as mitigating factors.


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3.3.2 Influences of impoundments on water chemistry

As with the thermal regime, impoundments may alter downstream water chemistry via two broad physical mechanisms: by increasing the retention time of water within the system and reducing its velocity (thus facilitating changes in chemical composition of the water); and by changing the timing, frequency, magnitude, duration and rate of change of releases to the downstream watercourse (Petts, 1986; Olden and Naiman, 2010).These changes affect the physico-chemistry of the water retained by the impoundment, and therefore also of any water either abstracted or released downstream. The processes driving such changes are described below.

The low velocity conditions generated by impoundments facilitate settlement of suspended solids (Petts, 1986) which may reduce the suspended solids concentration of any releases, depending on the type and position of the draw-off mechanism. Impoundments with surface releases can reduce downstream turbidity as a consequence of this settlement effect (unless there is a high algal content within the impounded water; algal biomass will tend to be concentrated in the surface layers of the water column (i.e. to the depth to which sunlight can penetrate) regardless of whether or not the reservoir is stratified). Impoundments with deep water releases can have high turbidity and suspended solids concentrations, particularly if they are used to remove accumulated sediment from behind the dam: so called scour-releases. The first release after an extended period of low flow can cause a marked reduction in dissolved oxygen concentration due to disturbance of the periphyton and organic debris on the streambed (Petts, 1986).

Electrical conductivity has been reported to exhibit lower variability downstream of impoundments in comparison with the un-impounded river (Soja and Wiejaczka, 2014; Petts, 1986), although this may be associated with the stability of the flow release from many reservoirs. Although Crisp (1977) reported no clear indication of changes in ionic flux as water passed through Cow Green reservoir, he does comment on the clear effect of impoundment in smoothing out the large short-term fluctuations of ionic content which are characteristic of the natural River Tees. Other studies have reported notable changes in conductivity and other water chemistry parameters as a consequence of impoundment (Brooker, 1981).

Soja and Wiejaczka (2014) identified a narrower range of conductivity in the Ropa River downstream of the Klimkowka reservoir in the Polish Carpathians, as well as a reduction in average conductivity following reservoir construction. However, the authors also acknowledged some uncertainty regarding the cause, with downstream flow stabilisation and storage of low conductivity meltwater within the reservoir (followed by gradual release during the year) being cited as possible contributing factors. There was reportedly no obvious effect of the reservoir on downstream pH (Soja and Wiejaczka, 2014). It should be noted that measurements of conductivity are temperature dependant (conductivity increases with increasing temperature) and therefore changes in the downstream thermal regime may also affect conductivity.

Characteristically stable conductivity was observed in the River Tees downstream of Cow Green reservoir in Northumberland (possibly associated with the stability of the release regime) (Petts, 1986). Downstream of Cow Green reservoir, annual potassium loadings have been reduced by 60% of normal but maxima are only 30% of those in the inflows, whilst calcium maxima have been reduced by 75% and annual minima increased by nearly 100% (Petts, 1986).


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In a similar way to effects on the thermal regime, of particular importance to downstream water chemistry is reservoir stratification. This can affect the quality of water released to the river downstream in a variety of ways depending on the time of year, the geographical setting of the reservoir, its physical characteristics (e.g. depth and surface area), and its operational regime (including the type and position of the release mechanism). Indeed, development of stable stratification conditions associated with a hypolimnial release mechanism is likely to be one of the major risk factors associated with adverse effects on downstream water chemistry.

In deep, stratified reservoirs, little mixing occurs below the thermocline during the summer (so there is no re-aeration of the hypolimnion via mixing) and since sunlight does not penetrate, there is no re-aeration via photosynthesis (Petts, 1986). This can lead to low oxygen conditions (anoxia) in the hypolimnion. Under such low oxygen concentrations anaerobic decomposition occurs, creating a low pH reducing environment which allows release of various substances e.g. hydrogen sulphide, carbon dioxide, iron and manganese from sediments (Hutchinson, 1975). Hypolimnia are also typically characterised by high concentrations of calcium, bicarbonate, silicate, ammonia, nitrogen and phosphate (Petts, 1986) which can also be due to sediment release. Thus, deep water hypolimnetic releases from stratified reservoirs can have low dissolved oxygen, high nutrient and high iron and manganese concentrations (these latter two can form a precipitate on the downstream river bed often mixed with decomposing algae).

Depending on their operation, reservoirs may also act as nutrient sinks (Petts, 1986). Toms et al. (1975), cited by Brooker (1981), concluded that in Grafham Water, a shallow reservoir in the UK receiving water pumped from a nearby river, there were reductions in the concentrations of the major plant nutrients NO3-N, PO4-P and dissolved silica and in calcium concentration, presumably as a consequence of utilisation by algae.

Ahearn et al. (2005) report that 17 reservoirs surveyed in central California as part of a 1979 USA Department of Water Resources study found significant nutrient increases associated with the onset of drought, but that it was the shallow reservoirs that were the most impacted. The same authors note that in a 2001 analysis of five major reservoirs in the USA, reservoirs in different regions processed nutrients differently, suggesting that site-specific factors may be particularly important with respect to nutrient cycling. Reservoir passage decreased NO3-N fluxes in the Rio Grande, increased NO3-N fluxes in the Lower Columbia, and had little effect on NO3-N in the Colorado.

The quality of hypolimnial water typically worsens until mixing occurs (possibly until the autumn overturn), indeed water with a higher solute concentration is typically denser, potentially exacerbating the stability of the stratification. This process can be exacerbated for reservoirs (as opposed to natural lakes) where summer drawdown, followed by vegetative colonisation, followed by re-submergence can lead to anoxic bottom water conditions (Petts, 1986).

In smaller and shallower lakes where complete mixing occurs throughout the year (and thus stable stratification does not develop), sufficiently high dissolved oxygen concentrations are maintained throughout the water column to allow aerobic decomposition to take place. This prevents the formation of anaerobic reducing conditions and the release of undesirable gases and ions from the bottom sediments. Such conditions can also be generated in deeper reservoirs via artificial de-stratification.


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Although surface releases are typically associated with low dissolved solids concentrations, in the warm epilimnial water of a stratified reservoir, phytoplankton can proliferate, releasing oxygen, and depleting the water of bicarbonate and nutrients. In eutrophic reservoirs, severe algal blooms may develop, and these can affect downstream water quality, particularly if there is a surface release from a stratified reservoir (or a release of any depth where there is continual mixing e.g. via artificial de-stratification). For example, Grabowska (2012) reports that a shallow eutrophic reservoir (Siemianówka Dam Reservoir (sic.)) leads to deterioration of water quality in the lowland river basin through changes in the phytoplankton composition and abundance. This was due to dominance of cyanobacteria. Water retention time, daily water outflow and TN:TP ratio were identified as the factors favouring cyanobacteria development in the reservoir with conditions in the river downstream also favouring development (meandering character and low flow velocities in the outflowing river, together with the method of operating the dam gates).

It has also been reported that other chemical changes may occur where water passes over high dams into a deep pool, or through hydroelectric turbines. In such cases, the entrainment of air has been reported to result in supersaturation of oxygen and nitrogen (Ebel (1969) and Beiningen and Ebel (1970) cited by Brooker, 1981), which may have local effects on biota. However, this issue is considered unlikely to be a major factor in the UK where the majority of dams are relatively small in a global context. This issue is therefore not considered further.

Brooker (1981) notes that the use of multi-level draw offs is now widely practiced to minimise the release of poor quality water and, in addition, that considerable effort has been directed at developing engineering procedures to promote de-stratification in reservoirs.

Lessard and Hayes (2003) found that water chemistry variables that are often changed downstream of deep release dams e.g. total phosphorus, conductivity and dissolved oxygen concentration, were not significantly altered by the small dams in their study (but the size of the small dams and the position of the release in each case is not reported).

Ahearn et al. (2005) note that although the effect of impoundment on dissolved silicon (DSi) has been given relatively little attention in the literature, results from those studies focussing on this topic have been very consistent; impoundments and lakes from Finland and Yugoslavia all tend to act as significant DSi sinks (Conley et al. 2000; Friedl et al. 2004; Humborg et al. 1997, 2000). The result in many cases seems to be downstream shifts in phytoplankton communities from siliceous to non-siliceous species (Humborg et al. 1997, 2000).

In summary, the major risk factors for adverse effects on downstream water chemistry appear to be:

Hypolimnial releases from deep reservoirs with long retention times, low exposure to wind-induced mixing and a small surface area to volume ratio (i.e. conditions which allow stable stratification to develop). In these circumstances the hypolimnion may become depleted of dissolved oxygen such that low pH, reducing conditions prevail at the sediment-water interface, allowing release of nutrients and dissolved metals such as iron and manganese.

If used infrequently, hypolimnial releases (even in the absence of stable stratification) may disturb settled organic matter and cause high downstream concentrations of suspended solids.


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Releases from eutrophic reservoirs may have high concentrations of algae and chlorophyll a, with low concentrations of dissolved nutrients (that have been utilised by the algae) and high pH. This situation may occur in epilimnial releases from eutrophic reservoirs with stable stratification, or in releases of any depth from small, shallow reservoirs and those that are continuously mixed.

3.3.2.1 Downstream propagation of water chemistry effects and recovery

Effects on downstream water chemistry depend not only on the limnological processes occurring within the reservoir (which depend on reservoir characteristics such as retention time, trophic status and exposure to wind-induced mixing), but also on the mechanism of the release and the characteristics of the downstream river and its catchment.

Studies from the global literature indicate that (in a similar way to thermal effects) impacts on downstream water quality can occur over long distances, particularly downstream of large dams. For example, although natural re-aeration can raise oxygen concentrations within a short distance downstream of a dam, depressed dissolved oxygen concentrations have been reported for 100 km below Hume Dam in Australia (Walker et al. 1979).

In low-energy downstream environments, anoxic hypolimnial releases can cause low dissolved oxygen concentrations for long distances downstream, although air drafts in the release port and turbulence in the tail-waters can often introduce enough oxygen to alleviate this effect. Such issues may be particularly acute when rivers receive hypolimnial releases from several impoundments (Petts, 1986). Brooker (1981) reports that reaeration of water discharged from Cherokee Dam in Tenessee Valley, USA was extremely slow and dissolved oxygen concentrations increased from less than 1 mg/l at the dam outlet, to only about 3 mg/l by 16 km downstream and about 6 mg/l by 70 km downstream. He notes that this contrasted sharply with water in the Little Tennessee River downstream of Calderwood Dam, USA where dissolved oxygen concentrations increased from 65 % saturation at the dam to 93 % saturation within 6 km.

Brooker (1981) cites Edwards and Crisp (1980) in concluding that in shallow, fast-flowing upland rivers, only local effects of hypolimnial releases on downstream dissolved oxygen concentrations would be expected, since the released water would be rapidly re-aerated via natural processes.

Petts (1986) concurs that downstream of many reservoirs in the UK, such as the River Derwent downstream of Ladybower reservoir in Derbyshire, downstream water chemistry effects associated with reservoir storage are limited to a short reach below the dam, and may only be detectable during summer stratification.

Downstream of Cow Green reservoir, intensive assimilation and aeration in the River Tees, together with the influence of tributaries causes water quality to approximate the original values within a relatively short distance downstream (Petts, 1986, Crisp, 1977). These effects are likely to be associated with both the well-mixed nature of the reservoir and the aerating effect of the Cauldron Snout waterfall downstream of the dam over which the River Tees drops about 40 m in 135 m (Armitage, 2006).

The effects of compensation and regulation releases from Caban Coch on pH and dissolved oxygen concentration were restricted to the River Elan (5 km downstream of the dam) and


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both parameters increased with increasing distance downstream of the dam (Ellery and Wilkins, 1999).

Mackie et al. (1983) noted that dissolved oxygen concentrations returned to reservoir inflow values within 1 km downstream of the large impoundment (12 m deep; 400 ha area) of Guelph Lake, Ontario (following discharge of anoxic waters from the hypolimnion).

Effects on dissolved oxygen concentration may be even less extensive downstream of small, well-mixed impoundments; Murphy (2011) reported no dissolved oxygen effects downstream of Altnahinch Dam, and Tosney (2013) noted minimal effects of scour releases on dissolved oxygen concentration downstream of reservoirs in Yorkshire, with any effects concentrated at a single site.

Alongside effects on dissolved oxygen concentration and pH, the longevity of effects on metal precipitates are also noted in several studies. Immediately downstream of Kielder Water in Northumbria and at several sites downstream (distance not noted), a bed deposit was reported comprising fine silt held together by a matrix of algae. When analysed it was found to contain 5% iron, 5% aluminium and 1% manganese by dry weight. During periods of low discharge in summer months the deposits were found to be up to 2 cm thick. During periods of compensation flow only the deposit was particularly widespread. When HEP generation was taking place the deposits frequently appeared to be moved downstream; they were found periodically at other sites along the river (distance not noted) Haile, James and Sear (1989).

Wightman et al. (1990) noted that the extent of historical downstream acidification below Llyn Brianne depended on the rate of release from the reservoir in relation to the natural tributary flows which would buffer and dilute the acidic water so reducing its effect. Monitoring of releases undertaken in 1985 for fisheries purposes showed a significant reduction in pH and increase in aluminium concentration down to the confluence with the Bran (~20 km downstream). Precipitates rich in manganese were observed blanketing the bed of the Tywi below Llyn Brianne down to the confluence with the Doethie (~3 km) (Wightman et al. 1990).

No experimental information was identified in relation to the downstream propagation of effects on BOD, solutes or nutrients, presumably due to their rapid assimilation by biota. Information on the downstream deposition of suspended solids and algal particles was also unclear – this is likely to be driven primarily by the quantity and dynamics of flow releases which drive the sizes of particles that are either deposited or remain in suspension.


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Table 3.2 Types of chemical modification downstream of impoundments

Parameter Influence and mechanism Risk factors Mitigation measures Sources

Water Framework Directive physico-chemical elements

Dissolved oxygen concentration

Concentration may be reduced if the release is made from the hypolimnion under conditions of stratification (since there is no mixing of the hypolimnion with the oxygenated epilimnion). This effect may occur downstream if there is high oxygen demand in the released water, even if the water is initially temporarily aerated by the action of being released. Mackie et al. (1983) report that discharge of anoxic waters through a bottom reservoir port resulted in a significant drop in oxygen concentrations immediately downstream, but that concentrations had recovered again within 1km downstream. Brooker (1981) cites Edwards and Crisp (1980) in concluding that in shallow, fast-flowing upland rivers, only local effects of hypolimnial releases on downstream dissolved oxygen concentrations would be expected, since the released water would be rapidly re-aerated via natural processes. However, Brooker (1981) also reports more extensive effects from some USA dams.

Hypolimnial releases. Stable stratification. Little re-aeration either via release mechanism or downstream watercourse characteristics (Certain), but, given the prevalence of studies demonstrating no significant effects this is considered likely to occur only in the UK for particular combinations of the above risk factors and to extend only for short distances downstream of impoundments, as rivers are relatively shallow

Turbulent release mechanism which encourages re-aeration. Turbulent downstream watercourse which encourages re-aeration. Surface or variable releases. Destratification.

Mackie et al. (1983) Brooker (1981) Petts (1986) Ellery and Wilkins (1999) note a decrease during regulation releases. Murphy (2011) reports no effect d/s of Altnahinch Tosney (2013) noted minimal effect of scour releases on DO in Yorkshire, any effects concentrated at a single site. Crisp (1977) notes that regulation has not appreciably altered the dissolved oxygen concentration of the River Tees d/s of Cow Green, but there is a waterfall shortly d/s of the dam. Edwards and Crisp (1980) cited by Brooker (1981) concluded that in fast, shallow upland rivers rapid equilibration was likely and only local impacts would be expected

Daytime supersaturation and a concomitant night time sag may occur in epilimnial releases from reservoirs with stable stratification and high algal concentrations, and/or surface releases from eutrophic reservoirs. Conceptually reasonable but few supporting sources identified.

Epilimnial releases. Eutrophic status. (Uncertain)

Variable depth releases. Catchment management to reduce nutrient inputs if eutrophication is anthropogenic

Petts (1986)

BOD Reduced concentration if the release is epilimnial and particulate matter has settled out and/or algae have utilised the majority of available organic matter. Conceptually reasonable but few supporting sources identified.

Epilimnial releases. Large impoundments with long residence time (Uncertain)

Variable depth releases Petts (1986)

Increased concentration if hypolimnial releases disturb and cause release of settled organic matter. Conceptually reasonable but few supporting sources identified.

Hypolimnial releases subject to intermittent use (Uncertain)

Variable depth releases. Mixing of water from various depths prior to release. Timing of scour releases to coincide with natural high flow or spill events

Petts (1986)

Ammonia Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification (since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may therefore be generated in the hypolimnion). Conceptually reasonable but few supporting sources identified.

Hypolimnial releases. Stable stratification (Uncertain)

Surface or variable releases. Destratification

Petts (1986)

pH pH may be low if the release is made from the hypolimnion under conditions of stable stratification. This effect may be exacerbated by stable stratification since there is no mixing of the hypolimnion with the oxygenated epilimnion. Lack of mixing can lead to low oxygen conditions (anoxia) in the hypolimnion which can result in anaerobic decomposition and low pH. Such an environment is often termed a “reducing environment” since conditions favour “reduction” as opposed to “oxidation” reactions. Wightman et al. (1990) reported exacerbation of low pH issues d/s of Llyn Brianne due to stratification

Hypolimnial releases. Stable stratification (Quite Certain)

Surface or variable releases. Destratification

Petts (1986) Wightman et al. (1990) Edwards and Crisp (1980) cited by Brooker (1981) note decrease in alkalinity Tosney (2013) notes a decrease during scour releases


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and hypolimnial draw-off. Ellery and Wilkins (1999) note a decrease during regulation releases

pH may be high if the release is made from the epilimnion and there is a high concentration of algae (NB releases of any depth could cause this in well-mixed eutrophic reservoirs). Conceptually reasonable but no supporting sources identified.

Epilimnial releases. Eutrophic status (Uncertain)

Variable depth releases. Catchment management to reduce nutrient inputs

Phosphorus Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification (since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may therefore be generated in the hypolimnion, causing release of nutrients as well as other substances). Release of nutrients from the sediment under anoxic conditions may be exacerbated where nutrient inputs to the reservoir have historically been high. If stratification occurs during the summer months, concentrations released during the summer months may be higher than would be expected in an unregulated catchment, leading to excessive algal growth downstream (depending on which nutrients are limiting to algal growth in the downstream catchment). Excessive nutrient releases from the sediment may be reduced or prevented by artificial destratification.

Hypolimnial releases during stable stratification. Sediments with high nutrient content and/or high nutrient concentrations in reservoir inflows. (Quite Certain)

Variable depth releases. Destratification. Catchment management to reduce nutrient inputs if the source of phosphorous is anthropogenic.

Petts (1986) Brooker (1981) Stanford and Hauer (1992) note an increase downstream of Hungry Horse Dam, Montana, which had a deep release until selective withdrawal was implemented in 1996. Ahearn et al. (2005) Zhong and Power (1996) note that nutrients are retained in the reservoir behind a high head dam in China with a deep release (Xinanjiang Dam) but this may be a consequence of storage of runoff events.

In contrast, concentrations released from the hypolimnion during the winter months when the reservoir is well-mixed may be lower than expected in an unregulated catchment due to storage of runoff events (which typically have high nutrient loading) within the reservoir. Epilimnetic releases under conditions of stable stratification tend to be nutrient poor (there is no mixing with the hypolimnion into which the nutrients are released from the sediment). This may be exacerbated if there is a high algal concentration which has depleted the surface waters of dissolved nutrients. The effects of stratification on nutrient release may therefore affect the timing of nutrient delivery to downstream reaches. The release of nutrients may be very variable between impoundments depending on site-specific factors.

Hypolimnial releases without stable stratification during high runoff events. Epilimnial releases during stable stratification (Quite Certain)

Variable depth releases. Artificial destratification. Reinstatement of high runoff events past the reservoir.

Petts (1986) Toms et al. (1975), cited by Brooker (1981) note a decrease d/s of shallow Grafham Water. Ahern et al. (2005) Zhong and Power (1996) note an increase in nutrients and phytoplankton below a high head surface release dam in China (Danjiangkou reservoir)

Temperature n/a discussed above n/a discussed above n/a discussed above n/a discussed above

Specific pollutants

Manganese Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification. Since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may be generated in the hypolimnion, allowing release of manganese from the sediments.

Hypolimnial releases. Stable stratification (Certain)

Surface or variable releases. De-stratification

Summers (undated a) Wightman et al. (1990) Gantzer et al. (2009) Gilvear et al. (2002) citing Bryant et al. (1997). Thomas (2005) Haile, James and Sear (1989) Brooker (1981) Petts (1986)

Iron Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification. Since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may be generated in the hypolimnion, allowing release of iron from the sediments.

Hypolimnial releases. Stable stratification. (Quite Certain)

Surface or variable releases. De-stratification.

Gantzer et al. (2009) Haile, James and Sear (1989) Brooker (1981) Petts (1986)

Other metals Aluminium

Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification. Since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may be generated in the hypolimnion, allowing release of metals from the sediments.

Hypolimnial releases. Stable stratification. Catchment acidification

Surface or variable releases. De-stratification

Summers (undated a) reports aluminium in precipitate d/s of Lubreoch Dam. Haile, James and Sear (1989) Wyke (1997) reports higher


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(Quite Certain) aluminium d/s of Llyn Brianne before liming began.

Pesticides / other organic compounds

No sources identified - - -

Other elements

Nitrate Concentration may be increased if the release is made from the hypolimnion under conditions of stable stratification (since there is no mixing of the hypolimnion with the oxygenated epilimnion, and reducing conditions may therefore be generated in the hypolimnion, causing release of nutrients as well as other substances). Release of nutrients from the sediment under anoxic conditions may be exacerbated where nutrient inputs to the reservoir have historically been high. If stratification occurs during the summer months, concentrations released during the summer months may be higher than would be expected in an unregulated catchment, leading to excessive algal growth downstream (depending on which nutrients are limiting to algal growth in the downstream catchment). Excessive nutrient releases from the sediment may be reduced or prevented by artificial destratification.

Hypolimnial releases during stable stratification. Sediments with high nutrient content and/or high nutrient concentrations in reservoir inflows. (Quite Certain)

Variable depth releases. Destratification. Catchment management to reduce nutrient inputs if the source of nitrate is anthropogenic.

Petts (1986) Brooker (1981) Stanford and Hauer (1992) note an increase downstream of Hungry Horse Dam, which had a deep release until selective withdrawal was implemented in 1996. Ahearn et al. (2005) Zhong and Power (1996) note that nutrients are retained in the reservoir behind a high head dam in China with a deep release (Xinanjiang Dam) but this may be a consequence of storage of runoff events.

In contrast, concentrations released from the hypolimnion during the winter months when the reservoir is well-mixed may be lower than expected in an unregulated catchment due to storage of runoff events (which typically have high nutrient loading) within the reservoir. Epilimnetic releases under conditions of stable stratification tend to be nutrient poor (there is no mixing with the hypolimnion into which the nutrients are released from the sediment). This may be exacerbated if there is a high algal concentration which has depleted the surface waters of dissolved nutrients. The effects of stratification on nutrient release may therefore affect the timing of nutrient delivery to downstream reaches. The release of nutrients may be very variable between impoundments depending on site-specific factors.

Hypolimnial releases without stable stratification during high runoff events. Epilimnial releases during stable stratification (Quite Certain)

Variable depth releases. Artificial destratification. Reinstatement of high runoff events past the reservoir

Petts (1986) Toms et al. (1975), cited by Brooker (1981) note a decrease d/s of shallow Grafham Water. Ahearn et al. (2005) Zhong and Power (1996) note an increase in nutrients and phytoplankton below a high head surface release dam in China (Danjiangkou reservoir).

Suspended solids Reduced concentration if particulate matter has settled out and/or algae have utilised the majority of available organic matter.

Epilimnial releases. Large impoundments with long residence time. (Quite Certain)

Sediment management plan Petts (1986) Crisp (1977) infers reduction in sediment movement past the dam following impoundment. Zhong and Power (1996) note a decrease below a high head dam in China

Increased concentration possible in epilimnial releases from reservoirs with stable stratification and high algal concentrations. Increased concentration possible from deep/scour releases. Wightman et al. (1990) reported increased sedimentation downstream of Llyn Brianne as a consequence of the hypolimnial draw-off. These are noted as likely to be important for up to 3 km but not likely to account for ecological paucity further downstream. Downstream sediment deposition may be driven by the quantity and dynamics of the flow release regime, as well as movement of different classes of sediment downstream of the impoundment. Capacity for re-suspension of material downstream may also be important.

Epilimnial releases. Eutrophic status. Deep/scour releases. (Quite Certain)

Variable depth releases. Catchment management to reduce nutrient inputs.

Petts (1986) Thomas (2005) Haile, James and Sear (1989) Wightman et al. (1990) Tosney (2013) increase during scour releases

Conductivity Electrical conductivity has been reported to be reduced and to exhibit lower variability downstream of impoundments in comparison with the un-impounded river (Soja and Wiejaczka (2014), Petts (1986), Crisp (1977)), although this may be associated with the stability of the flow release from many reservoirs. Unknown if this effect differs by, for example, type or size of impoundment.

Impoundment. (Quite Certain) Increased conductivity

Variable depth releases. Artificial destratification.

Soja and Wiejaczka (2014). Crisp (1977). Petts (1986).


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Conductivity may be increased by releases from sediment associated with stable stratification and hypolimnial releases. However, concentrations released from the hypolimnion during the winter months when the reservoir is well-mixed may be lower than expected in an unregulated catchment due to storage of runoff events (which typically have high conductivity) within the reservoir.

possibly exacerbated by hypolimnial releases and stable stratification. Reduced conductivity relative to upstream catchment possible when runoff events are stored in the reservoir.

Reinstatement of high runoff events past the reservoir

Algae / chlorophyll a Increased concentration possible in epilimnial releases from reservoirs with stable stratification and high algal concentrations, and/or releases from eutrophic reservoirs, especially problematic if released into slow-flowing downstream reaches.

Epilimnial releases. Eutrophic status. (Quite Certain)

Variable depth releases. Catchment management to reduce nutrient inputs if the source of excessive nutrients is anthropogenic.

Haile, James and Sear (1989) note that river deposits d/s of Kielder composed of fine silt held together by a matrix of algae with high metal (Mn, Fe, Al) content. Associated with hypolimnial release and exacerbated by compensation discharge only. Petts (1986)

Solutes Ca, K, Mg, Na Concentrations may be lower in surface releases.

Epilimnial releases. Surface releases (Quite Certain)

Variable depth releases Artificial destratification Reinstatement of high runoff events past the reservoir

Petts (1986) citing Wetzel (1975) and Gore (1980 Crisp (1977) Ahearn (2005)

Concentrations may be higher in hypolimnial releases under conditions of stable stratification. Relatively high Ca/Na ratios, especially during the autumn and winter months, may occur in hypolimnial releases (Wetzel, 1975; Gore, 1980, cited by Petts, 1986). Conceptually reasonable but few supporting sources identified.

Hypolimnial releases. Stable stratification (Quite Certain)

Variable depth releases Artificial destratification

Petts (1986)

Dissolved silicon Ahearn (2005) notes that there is limited literature on the effect of impoundment on dissolved silicon but that the results of available studies are very consistent; that impoundments and lakes tend to act as significant dissolved silicon sinks. The result in many cases seems to be downstream shifts in phytoplankton communities from siliceous to non-siliceous species (Humborg et al. 1997, 2000).

Impoundment (Quite Certain)

- Ahearn (2005) citing: (Conley et al. 2000; Friedl et al. 2004; Humborg et al. 1997, 2000)


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3.4 Influences of impoundment thermal and water chemistry effects on ecology

Those effects of impoundments on ecology mediated via changes in flow regime and sediment dynamics have been extensively reported elsewhere (e.g. Petts, 1984; Higgs and Petts, 1988; Bunn and Arthington, 2002; SNIFFER WFD21D) but are not of primary interest to this study which is focused on the effects on water temperature and water chemistry, and associated effects on ecology.

However, as noted by Olden and Naiman (2010), in general it remains difficult to isolate quantitatively the relative contributions of water temperature (and other water quality variables influenced by dams) and water flow to observed biotic responses in riverine ecosystems (Bednarek and Hart, 2005). This finding is supported by comprehensive reviews of the literature by Haxton and Findlay (2008) and Gillespie et al. (2014), both of which highlight the limited number of studies that quantify the water chemistry and temperature effects of dam-induced change on downstream ecological responses (as distinct from effects arising from changes to flow regime and sediment dynamics). Notwithstanding the overwhelming consensus that riverine organisms, in situ, respond to a combination of physical and chemical changes downstream of impoundments, the following narrative describes the specific effects of water temperature and chemistry from the global scientific literature, highlighting the UK situation.

Water temperature has direct effects on the metabolic rates, physiology and life history traits of aquatic species, helping to determine rates of important ecological processes such as nutrient cycling and productivity, and indirectly mediating biotic interactions (Olden and Naiman, 2010 citing: Magnuson et al. 1979; Petts, 1986; Poole and Berman, 2001; Caissie, 2006; Webb et al. 2008).

Olden and Naiman (2010) highlight several fundamental ways in which the natural thermal regime of a river provides cues and thresholds that trigger and shape important ecological events:

the accumulation of daily maximum temperatures above a critical threshold has been shown to be a fundamental parameter in shaping the distribution and condition of many aquatic species (Armour, 1991);

development schedules for freshwater fish and insects respond to the summation of thermal units (i.e. the accumulation of daily temperatures above some threshold) as well as absolute temperatures, and fish species have both chronic and acute temperature thresholds for survival, growth and reproduction (Vannote and Sweeney, 1980; Coutant, 1987);

natural thermal cues stimulate fish migration, spawning and egg hatching, and directly influence egg survivorship and developmental time (Wootton, 1990); and

changes in temperature can influence the type and prevalence of diseases affecting fish.

This report has shown that the effects of impoundments on downstream thermal regime and water chemistry vary depending on the time of year, the geographical setting of the reservoir, its physical characteristics (e.g. depth and surface area), and its operational regime (including the type and position of the release mechanism). The knock-on effects of such changes on ecological thresholds and cues are therefore also dependent on these factors, and can differ widely depending on site-specific conditions.


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This section of the report takes the processes and risk factors, identified in the preceding sections of this report that are associated with water temperature and chemistry effect of large impoundments, and describes their documented effects on riverine biota. Components of riverine biota are described in turn with respect to the biological quality elements of WFD water body classification: macrophtyes; phytobenthos (algae and diatoms); macro-invertebrates; and fish.

3.4.1 Macrophytes and phytobenthos

There is relatively little information on the effects of large impoundments on downstream aquatic macrophytes and phytobenthos, compared to macro-invertebrates and fish. Macrophytes and algae typically increase in abundance and richness downstream of large impoundments and this is mostly attributed to the more stable hydraulic conditions imposed by impoundments, rather than any explicit reference to water temperature and chemistry effects. Those studies that do reference the effects of water temperature and chemistry do not do so in isolation from the interacting effects of hydraulics and morphology.

Jackson et al. (2007) note that algae is frequently abundant below impoundments (citing Inverarity et al. 1983; Allan, 1995; and Penaz et al. 1999 as examples of this phenomenon), principally in circumstances where the absence of high flows reduces bed scour. The increased growth of algae in the Lyon has been a notable feature of the river since the construction of the dams (Summers, 2000).

Water temperature effects from large impoundments have been described only for epilithic algae; no such effects were described for macrophytes. Two studies on rivers in Montana suggested apparent contrasting effects of water temperature that increased the coverage of epilithic algal mats downstream of large reservoirs. One study suggested that slightly higher than ambient water temperatures downstream of a large impoundment in winter increased algal growth (Stanford et al. 1988 cited in Stanford and Hauer, 1992). The other study suggested that cooler than ambient water temperatures downstream of another impoundment promoted algal growth (Gore, 1980). It is not clear if the latter study referred to summer cooling, but this can be reasonably assumed given the temperate climate and characteristics of the impoundment. Summer cooling and winter warming from hypolimnetic releases in these settings could have resulted in a more constant temperature regime centred on the algae’s thermal optimum.

Increased nutrients released from the bottom of large impoundments of eutrophic reservoirs increases the coverage and taxon richness of submerged macrophytes as well as the coverage of algae. Algae derived from nutrient-rich reservoirs can also establish a phytobenthos characteristic of a eutrophic river (Petts, 1986).

Table 3.3 Biotic responses to thermal and chemical effects downstream of impoundments – macrophytes and phytobenthos

Water chemistry and thermal variables affected

Biotic responses Sources


- -

Reduced magnitude of seasonal temperature variation (increased seasonal

- -


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constancy)


Increase in density of epilithic algal mats due to colder water temperatures in a prairie river in Montana

Gore (1980)


Increase in the periphytic biofilm with warmer winter water temperatures from Hungry Horse Reservoir, Montana

Stanford et al. (1988) Cited in Stanford and Hauer (1992)


- -


- -

Hypolimnial releases from stratified reservoirs with low dissolved oxygen, low pH, high concentrations of nutrients, high concentrations of dissolved metals

Hypolimnetic releases of high concentrations of phosphate and nitrate caused increased richness and coverage of submerged macrophytes

Autochthonous primary production in regulated rivers may be enhanced due to regular releases with low turbidity and stabilisation of the substrate

Hypolimnetic releases of high concentrations of phosphate and nitrate caused an increase of periphytic and perilithic diatoms and algae

Benitez-Mora and Camargo (2014)

(Petts, 1986)

Benitez-Mora and Camargo (2014); Marcus (1980)

Infrequently used hypolimnial releases with high concentrations of suspended solids.

- -

High concentrations of algae and chlorophyll a, with low concentrations of dissolved nutrients (that have been utilised by the algae) and high pH.

Algae derived from upstream impoundments (especially via epilimnial releases) have been found to proliferate in slow-flowing nutrient-rich regulated rivers, and impoundments providing considerable amounts of organic matter, alone, or together with inorganic nutrients can stimulate a downstream phytobenthos that is normally associated with nutrient enriched streams

(Petts, 1986)


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3.4.2 Macro-invertebrates

Large impoundments generally have the greatest effect on fauna as a result of hypolimnetic releases of water in continental and warm temperate climates (Saltveit et al. 1987; Brittain and Saltveit, 1989). Table 3.5 gives the full details of the literature review on this subject and the following narrative provides a summary, focussing on the UK situation.

Insects are most negatively affected by water temperature changes downstream of large impoundments because they rely on different temperature cues at different times of year to progress successfully through their life cycles. Molluscs and crustaceans are least affected and often increase in abundance in the absence of competition from thermally-sensitive insects and other conditions that become favourable below impoundments, such as increased algae for food and decreased flood disturbance (Gore, 1980; Vinson, 2001; Armitage, 2010). In general, temperature changes downstream of large impoundments result in reduced taxon richness and increased density of certain species (Stanford and Hauer, 1992).

Thermal constancy from deep water releases can cause non-lethal effects on aquatic insects by disrupting the timing of life cycles through reduced fecundity, reduced mating success, resulting in smaller and extinct populations (Sweeney, 1978).

Summer cooling from hypolimnetic releases can reduce growth rates of insect larvae, delaying emergence, decoupling male and female emergence and even preventing adult emergence, with dramatic consequences to populations of, especially, Ephemeroptera, Plecoptera and Trichoptera (Vinson, 2001). Summer temperature reductions of as little as 3oC below ambient have been linked to these effects in Plecoptera (Raddum, 1978).

Winter warming from hypolimnetic releases can increase growth rates of insect larvae causing early emergence, sometimes into lethally cold air temperatures (Ward and Stanford, 1979; Rader and Ward, 1988); Vinson, 2001). These conditions can also prevent the hatching of species from eggs that require near zero temperatures to stimulate hatching. Consistent with summer cooling, winter warming of the order of 2-3oC has been attributed to negative effects in insects below large impoundments in continental North America (Vinson, 2001).

Summer warming of water from epilimnial releases or from wide, shallow reservoirs has not been reported as having negative impacts on downstream invertebrates.

The effects of summer cooling from hypolimnetic releases are considered to have the most negative impacts in downstream invertebrates (Stanford and Hauer, 1992), especially in temperate climates, like the UK, where no invertebrates need a cold shock in their development or have a facultative winter diapause. In these climates, the negative effects of summer cooling from hypolimnetic releases present the main thermal risk for riverine insects. There is less information available on the effects of changes in water chemistry on invertebrates downstream of large impoundments. Increases in the concentration of nutrients and higher Ca/Na ratios in hypolimnetic releases from eutrophic reservoirs can increase overall invertebrate density and biomass, especially of molluscs, beetles, dipterans


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and amphipods (Gore, 1980). It can concomitantly cause a decrease in taxon richness and the abundance of Plecoptera, Trichoptera and some Ephemeroptera species (Gore, 1980; Benitez-Mora and Camargo, 2014). This illustrates the potential interacting effects of thermal alteration, increased nutrients and increased algae downstream of large, nutrient rich impounded water bodies, creating conditions that are favourable to molluscs and amphipods and unfavourable to Epemeroptera, Plecoptera and Trichoptera. The effects of impoundments on freshwater pearl mussels Margaritifera margaritifera have been reported due to hydraulic and morphological changes, but not due directly to water temperature and chemistry effects (Addy et al. 2012).

In summary and with focus on UK reservoirs:

Deep, eutrophic reservoirs present the greatest risk to downstream invertebrates as a result of thermal and chemical alterations;

In non-eutrophic reservoirs, the greatest risk to downstream invertebrates is from cooler summer temperatures disrupting the development and life cycles of insects. Winter warming is considered not to be such a large risk in the temperate climate of the UK, but this must be assessed on an individual site basis and adaptively managed.

The negative effects of water chemistry changes downstream of deep, non-eutrophic reservoirs are not well documented and whilst this is not considered to be a major risk to downstream invertebrates, it must be assessed on a site-specific basis and adaptively managed. Where effects have been found they have tended to be extremely localised to the area immediately below the impoundment, such as fine sediment, iron and manganese deposits affecting macro-invertebrates just below the impoundment of Rutland Water (Chris Extence, Environment Agency, pers. comm.). As part of this assessment, environment agencies must consider whether highly localised impacts from poor water quality affect the overall status of water bodies.

Whilst this review has attempted to isolate the relative impacts of altered water temperature and chemistry, the reality is that in all cases reviewed across the world, invertebrates respond to the combination of water quality, hydraulics and morphology changes below impoundments. In the UK, an investigation of ecological response downstream of Loch Lyon, Scotland, Jackson et al. (2007) found that individual flow or thermal metrics did not exert an overriding influence on invertebrate community structure, rather, whole regime changes were implicated in the impoverishment of invertebrate communities below the dam.


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Table 3.4 Biotic responses to thermal and chemical effects downstream of impoundments – macro-invertebrates

Water chemistry and thermal variables affected Biotic responses Sources


Seven-fold increase in downstream drift density of invertebrates in River Wye, also linked to reduced discharge variation. Slater (1978) Unpublished.

Thermal variance maximises the number of taxa that can coexist. Vannote et al. (1980) Cited in Vinson (2001)

Constant conditions caused by deep hypolimnetic releases cause a reduction in biodiversity, but biomass of certain species can be very high.

Stanford and Hauer (1992)

The absence of thermal variance can cause increased competition among stream insects. Thermal constancy can disrupt normal growth and emergence patterns in mayflies, causing reduced fecundity, reduced mating success, leading to smaller and extinct populations.

Sweeney (1978)

Non-insects, such as amphipods that do not have critical maximum and minimum temperatures to complete their life-cycles can increase in abundance and dominate communities downstream of hypolimnetic release dams. Armitage (2010) reported increased abundance of Gammaridae immediately downstream of Cow Green Reservoir – but not explicitly related to reduced temperature variation

Vinson (2001); Angradi and Kubly (1993) Cited in Vinson (2001) Armitage (2010)

Favours taxa inhabiting constant temperature springs Vinson (2001)


Thermal variance maximises the number of taxa that can coexist. Vannote et al. (1980) Cited in Vinson (2001).

Constant conditions caused by deep hypolimnetic releases cause a reduction in biodiversity, but biomass of certain species can be very high.


The absence of thermal variance can cause increased competition among stream insects. Thermal constancy can disrupt normal growth and emergence patterns in mayflies, causing reduced fecundity, reduced mating success, leading to smaller and extinct populations.

Sweeney (1978)

Non-insects, such as amphipods that do not have critical maximum and minimum temperatures to complete their life-cycles can increase in abundance and dominate communities downstream of hypolimnetic release dams. Armitage (2010) reported increased abundance of Gammaridae immediately downstream of Cow Green Reservoir – but this was not explicitly related to reduced temperature variation.

Vinson (2001); Angradi and Kubly (1993) Cited in Vinson (2001). Armitage (2010)

Favours taxa inhabiting constant temperature spring outflows Stanford and Hauer (1992); Vinson (2001)


Seven-fold increase in downstream drift density of invertebrates in River Wye, also linked to reduced discharge variation. Slater (1978) Unpublished.

Reduced growth rate leading to non-maturation and non-emergence of insects. Ward and Stanford (1979); Rader and Ward (1988)

Persistence of endogenous invertebrates leading to dominance of molluscs because of both loss of temperature cues for insect development and increase in Ca/Na ratio in hypolimnetic releases.

Gore (1980); Petts (1984)

Advantageous to cold water stenotherms. Cold water stenothermic chironomids, Heterotrissocladiussp., Macropelopia sp. and Prodiamesia sp. were abundant below a Norwegian dam with lower summer water temperatures and absent above

Cereghino and Lavandier (1998) Saltveit, Bremnes and Brittain (1994)

Reduced average number of taxa of Ephemeroptera, but mean density is often increased. Brittain and Slatveit (1989)

Summer temperature reduction of 3oC caused increased stonefly abundance but decreased biomass. Raddum (1978)

Certain taxa need a specific maximum summer temperature to complete development. Lack of stoneflies downstream of Hungry Horse Reservoir, Montana was caused by a lack of attainment of a specific maximum summer temperature.

Ward and Stanford (1979)

Reduced summer water temperature was associated with a complete failure of a summer generation of Baetis tricaudatusin Red Creek.

Pearson et al. (1968) Cited in Vinson (2001)

Cold shock of summer hypolimnetic releases has been suggested as more detrimental to ecology in downstream rivers than winter warming.


Virtual elimination of Ephemeroptera, Plecoptera and Trichoptera in the Colorado River downstream of Glen Canyon Dam attributes to cold water releases preventing these insects from completing their life-cycles to emergence

Stevens et al. (1997) Cited in Olden and Naiman (2010)

In warm climates, like Spain, summer cooling by hypolimnetic releases might be favourable to benthic communities. Casado et al. (1989)


Premature emergence of insects and exposure to lethal cold air temperatures. Unbroken winter diapause. Elimination of winter diapause causing premature emergence and exposure to lethal cold air temperatures.

Ward and Stanford (1979); Rader and Ward (1988); Vinson (2001)

Persistence of endogenous invertebrates leading to dominance of molluscs because of both loss of temperature cues for insect development and increase in Ca/Na ratio in hypolimnetic releases.

Gore (1980)


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Water chemistry and thermal variables affected Biotic responses Sources

Advantageous to cold water stenotherms if only a slight increase in temperature towards thermal optima Cereghino and Lavandier (1998); Petts (1984).

Reduced average number of taxa of Ephemeroptera, but mean density is often increased Brittain and Slatveit (1989)

Winter increase of 1-2 oC caused increased stonefly abundance but decreased biomass Raddum (1978)

Winter increase in minimum temperatures of 2-3 oC eliminated many Ephemeroptera, Plecoptera and Trichoptera species after

dam closure. Suggested mechanisms were loss of physiological developmental cues, increased growth, earlier emergence, loss of diapause breaks.

Vinson (2001)

Experimental increases of winter temperature of +9.5 oC and +15

oC caused earlier emergence and smaller adults of Ephemerella

subvaria. Warmer temperature initiated the development of adult tissue in smaller larvae resulting in smaller adults emerging earlier than in ambient stream temperatures.

Vannote and Sweeney (1980)

Increases in winter temperature from hypolimnetic releases is suggested as less of an ecological impact in downstream rivers than summer cold shock



Favours warm water specialists Petts (1984)


Persistence of endogenous invertebrates leading to dominance of molluscs. Gore (1980)

Slower rate of warming in summer can reduce fecundity of emerging adult insects, exaggerate the separation of male and female emergence, prolong emergence time (increasing susceptibility to trout predation) and reduce larval growth rates so that emergence of the next generation occurs later when air temperatures are lethally cold.

Vinson (2001) and various references cites within.


Increase in Ca/Na ratio in cold hypolimnetic releases can favour the persistence and dominance of molluscs when combined with increased filamentous algal growth Hypolimnetic releases of high concentrations of phosphate and nitrate caused an increase in the abundance of molluscs, beetles, dipterans and amphipods; an increase in overall invertebrate density and biomass; and an increase in the abundance of scrapers and collector-gatherers. It also caused a decrease in taxon richness and the abundance of Plecoptera, Trichoptera and some Ephemeroptera species. There is anecdotal evidence of oxidised dissolved metals precipitating out of solution immediately downstream of large impoundments and causing negative effects on macroinvertebrates by smothering the river bed. Iron deposits in the River Gwash immediately downstream of Rutland Water have negatively affected macroinvertebrates.

Gore (1980) Benitez-Mora and Camargo (2014) Chris Extence, Environment Agency pers. comm.


Scour releases of hypolimnial water from Rutland Water releasing plugs of sediment has negatively affected macroinvertebrates in the River Gwash

Chris Extence, Environment Agency pers. comm.

High concentrations of algae and chlorophyll a, with low concentrations of dissolved nutrients (that have been utilised by the algae) and high pH.

Increase in molluscs, Gammarus pulex, Hydropsychidae, Serratella ignita and Hydroptilidae downstream of Cow Green Reservoir has been attributed in part to increased fine suspended organic material including algal fragments.

Armitage (2006)


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3.4.3 Fishes

Fish biology and life histories are genetically adapted to the environmental regimes typical of their natural habitat range (e.g. Verspoor et al. 2007), including water temperature (Jensen, et al. 2008). Water temperature is particularly important for fishes because they are poikilothermic (ectothermic) and therefore all their metabolic processes (e.g. egg incubation, digestion, growth, swimming capacity and many others) are directly influenced by temperature (e.g. Brett, 1956; Wootton, 1990). Water chemistry also exerts effects on fishes, mainly through physiological influences directly on them and on ecosystems components on which they depend, such as prey and predators. However, although water temperature is demonstrated to affect predator-prey interactions of fishes, too little is known about it to inform the management of flows (Nislow and Armstrong, 2012).

The responses of fishes to downstream environmental changes can be categorised into effects at individual or population / community levels. Individual level comprises physiological and behavioural responses. The former are comparatively well described and the principles apply across all species, with variation according to their specific physiology and ecological adaptations. Therefore, they are comparatively predictable, although are subject to the confounding effects of other habitat factors (Flodmark et al. 2004). As thermal or water chemistry regimes deviate from normal, individual fish adapt and regulate until behavioural responses make them move away, or physiological responses can no longer cope and life history impairment or death results. Population dynamics, such as effects of competition or ecosystem productivity and physical and hydraulic habitat can significantly modify the behavioural responses of individuals. Local channel structure and hydraulic features can modify the ability to withstand regime changes. For example, trout seek out thermal refuges in deep pools (Elliott, 2000) and adult salmon have been observed to seek out thermal refuges on a diurnal basis (Moore et al. 2012). Both behaviours are modified by flow and habitat features. Such confounding effects are important and complicate greatly the isolation and prediction of the specific impacts of temperature and water chemistry (Thorstad et al. 2008; Olden and Naiman, 2010; Nislow and Armstrong, 2012; Milner et al. 2012).

Many of the observations in Table 3.5 are from studies in non-regulated rivers or in experimental conditions, but nevertheless represent the proximate impacts of variable (temperature or water chemistry) change. Most of the data in the table refer to salmonids because they have been more intensively studied (Cowx et al. 2012). The main difference from the other major fish group, the coarse fishes and eel, are due to the general preference for cooler water of the salmonids. Thermal tolerance is higher in non-salmonids (Figure 3.3) and generally non-salmonid riverine fish demonstrate physiological tolerances in keeping with their adaptations to river zones with poorer water quality, for example warmer water and lower dissolved oxygen concentration (Davies 1975).


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Figure 3.3 Thermal tolerance ranges (min-max, oC) for reproduction in some common British fish,

showing the transition from cold water species - brown trout (Salmo trutta), Atlantic salmon (Salmo salar) and grayling (Thymallus thymallus); through cool water species – perch (Perca fluviatilis), pike (Esox

lucius), roach (Rutilus rutilus) and minnow (Phoxinus phoxinus); to warm water species – bleak (Alburnus alburnus), common bream (Abramis brama), chub (Leuciscus cephalus), silver bream (Blicca

bjoerkna) and tench (Tinca tinca). Adapted from Webb and Walsh (2004)

Table 3.5 distinguishes between seasons (summer and winter); but seasonal effectscan cancel out when their impacts extend over long developmental stages that span the switch from winter warming to summer cooling. Salmonid egg development lasts from autumn to spring; covering periods when hypolimnetic discharges can warm and cool downstream water alternately.

In contrast to effects on individual life stages (eggs, juveniles, adults), the overall population and community responses are less frequently studied and evidently more variable because they depend on the processes of population dynamics and community ecology, making generalised predictions extremely difficult (Rose, 2000; Poff and Zimmerman, 2010). Such higher level organisational and wider spatial scale responses are ultimately the most important in evaluating the fisheries impacts of downstream discharges. Therefore great care is required when evaluating the full implications of individual life stage impacts on fish populations and fisheries.

Overall downstream effects on population scale metrics, such as fish production, are sometimes conflicting. Crisp et al. (1983) reported an increase in trout and bullhead production downstream of Cow Green in spite of cooler water. Invertebrate abundance increased, but rather than a large increase in growth (which was constrained by the temperature), they found that the improved trophic conditions were taken advantage of by increased recruitment of younger age classes, which, being smaller, have a higher unit production rate. In a review of Canadian regulated rivers Burt and Mundie (1986) found that 76% had reduced fish populations downstream. They contended that most of the impacts were directly flow-related, but they suggested that 17% of the reductions might also have involved water temperature.

0 5 10 15 20 25 30

Brown trout

Atlantic salmon

Grayling

Perch

Pike

Roach

Minnow

Bleak

Common bream

Chub

Silver bream

Tench

Water temperature, min-max range (oC)


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In extreme cases, fish species can be eliminated below impoundments due to such effects as spawning failure or thermal regimes outside the optimal range for particular life history stages. In other scenarios, a regulated thermal regime can produce water temperatures within or near the optimum range, whilst lower summer water temperatures may reduce predator and competitive fish species (Petts, 1986).

Re-wetting of channels downstream of dams is not a guarantee of fish production and caution should be applied to the projected increase from standard hydraulic based habitat models, irrespective of their intuitive appeal and the notions of environmental flows. Whilst it appears to be obvious that some water will always be better than none, the final fisheries outcomes are not easily predictable and such practices need to be part of an adaptive management strategy (Bradford et al. 2011).

In some cases, hypolimnetic releases of cold water may provide unique and highly desirable fishing opportunities for trout or salmon in regions that could otherwise not support cold water species (Olden and Naiman 2010). However, this is unlikely to be the case in the UK where the majority of rivers regulated by high head dams would naturally have supported cold water species.

Water chemistry effects on fishes specifically resulting from dam releases are less well described than effects of flow and temperature. There is a huge literature on water chemistry impacts on fishes; some standards are given in Alabaster and Lloyd (1982) and more recent literature gives additional importance to the sub-lethal effects of organic chemicals of many kinds (e.g. Moore et al. 2007; Sumpter, 2009). The release of water under pressure can lead to super-saturation of gases resulting in Gas Bubble Disease (GBD), and this is reported to occur downstream of high head reservoirs. Most literature on GBD is from North America and Scandinavia and it is not reported as a routine problem in the British Isles. Salmonids are more susceptible to GBD than percids and cyprinids (Heggberget, 1984). Fish are thought to be unable to detect super-saturated gasses, but its importance varies with depth, because deeper swimming relieves the symptoms. (Johnson et al. 2010). However due to their comparatively modest size no GBD problems have been reported in British reservoirs (Brooker 1981) and is not regarded as an important topic for this review.

Fish homing depends on suitable water chemistry. When migratory anadromous species home to natal spawning areas they are believed to respond to olfactory cues which might include pheromones from kin and chemicals from the geochemistry of the local sub-catchment. It can be suggested that if the changes in water chemistry from dam releases are sufficiently large they might interfere with homing. There is little direct information on this process but there is some evidence indicating differential effects of released water and natural spates. Salmon migrating up the River Tyne were stimulated to move up from the estuary by both releases from Kielder reservoir and natural spates, but the migration in response to the natural spates was always greater (Archer et al. 2008; Bendall et al. 2012),

Intra-gravel conditions are also important. Hyporheic water chemistry (principally dissolved oxygen), was impaired in regulated rivers in Norway, making them less suitable for brown trout egg development (Calles et al. 2007). The impacts were worse in a river with no specified minimum flow compared with having a minimum flow.

The downstream changes in water temperature, chemistry (and other related habitat) conditions can offer some benefit for fish production through stabilised flows for example (Nislow and Armstrong, 2012); but their spatial extent is dependent on the local hydrology,


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dilution and mass-transfer dynamics (Brooker, 1981). They can also affect other biota, making it necessary to evaluate the consequences in an overall ecosystem context (Milner et al. 2012).

At Kielder Water in Northumbria, Haile, James and Sear (1989) report a winter warming thermal effect in the absence of winter stratification that caused adverse effects on salmonid egg hatching – too early due to warmer winter temperatures, followed by a delay in the spring temperature increase that will initiate feeding (particularly salmon as is 7oC rather than 4oC for trout), thereby causing starvation of alevins).


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Table 3.5 Biotic responses to thermal and chemical effects downstream of impoundments – fishes




No direct evidence on diurnal range known, but Elliott (1975) concluded no effect of daily temperature variation about means. Reduced range in a reservoir discharge could offer some minor localised benefit.

Edwards et al. (1979) Crisp (1977) Elliott (1975)

NB evidence from HEP hydro-peaking experimental simulation studies suggests a similar principle – the less unnatural variation the better for salmonid production.

Flodmark et al. (2004)


Using temperature-growth model simulation and observation data, trout growth predicted to increase with reducing water temperature monthly range, This was for streams of different type (e.g. groundwater-fed vs spate) as might arise with hypolimnetic releases. Note the strength of effect depends where the range lies within the overall tolerance limits.

Edwards et al. (1979)


Fish feeding and growth decrease with temperature below optimum temperatures, e.g. for G trout Topt=13.1-14.1oC (diet type dependent), for salmon

Topt=15.9oC. NB affected by acclimations, thus: trout growth range 4.9-18.0

oC at Taccl of 4.0

oC; range 6.3-19.0

oC at Taccl =19.0

oC

Elliott and Elliot (2010) Solomon and Lightfoot (2008)

Growth in salmonids experiencing hypolimnetic cool water releases will be less than in unmodified equivalent sites (See Tywi, S. Wales and Wye {Milner, unpublished]). The effects could be an increase in mean smolt age and fewer smolts due to increased FW residence time and resulting mortality. NB this runs counter to the current national climate-related long term reduction in mean smolt age that has been tentatively attributed to the on average increase in water temperatures in UK rivers; although recently (since about 2003/4) this pattern has been apparently reversed (Cefas/EA, 2014). Juvenile growth and mean smolt age of salmon and trout reduced by cold reservoir releases in River Surna, Norway. Similarly, in the Alta, salmon growth was reduced in cooler water, but over the full season the effects of warming and cooling almost cancelled out, resulting in small growth changes.

Elliott, 2000; Jensen et al. (1998) Cefas/EA (2014) Saltveit (1990) Bendall et al. (2012); Gowans et al. (1999)

Therefore cooling releases are potential means of climate adaptation, but note the comparatively small areas affected downstream of most dams in the British Isles.

Rustabakken et al. (2004) Moore et al. (2012)

Swimming ability, duration and secondary effects such as the ability to pass obstacles are related to temperature. Rates of passage observed to be lower at lower temperatures, and movement stopped on the River Eden at temperatures <5.5 to 8.5

oC; but there is usually a strong interaction with flow. The

temperature effect is also affected by the presence and difficulty of obstacles to passage, such as weirs and rapids. E.g. adult sea trout can pass a weir at 8

oC, but not at 6

oC. Salmon stop ascent of Pitlochry fish pass at 5.5

oC, and rapids on Cassley of varying difficulty at temperatures of 7.2-11.1

oC.

Solomon and Lightfoot (2008) Gibbins et al. (2001) Solomon and Sambrook (2004)

Salmonids (adults and juveniles) in rivers seek out cooler water when thermally challenged. They are therefore less thermally challenged in cooler water of reservoir releases in summer months. Juveniles observed to be more easily displaced by water velocity at lower temps; e.g. displacement velocities 13.5-14.6cm.s

-1 at 7-9

oC and 32-34cm.s

-1 at 13-

18oC. Thus on Tyne re Kielder releases the effects of flow increase greatest on small 0+ salmon (April) vs summer releases (July) when fish displaced less.

River entry of adult salmon is shown to be inhibited at high temperatures, e.g. > 20.0oC but any cooling effect from dams to counter this is unlikely to extend to

estuaries due to re-warming. Water temperature is, along with photoperiod, an important stimulus for smolting, (salmon more that sea trout) needing to exceed thresholds probably in the spring just before smolting rather than prolonged durations, e.g. 10

oC reported in several rivers, but range of 5.8-11.2

oC in the Imsa, Norway.

McCormick et al. (1998)

In temperate climates, decrease in summer water temperatures from hypolimnetic releases of water are beneficial to salmonids. Casado et al. (1989) and Cadwallader (1978); Cowx, Young and Booth (1987); Crisp, Mann and Cubby (1983) referenced within


Egg incubation times decreased by higher winter temps; but incubation effects can be wholly or partly offset at end of incubation periods by lower spring time temps, depending when the switch (warming to cooling) occurs (e.g. on the River Wye, Milner unpublished). Fish feeding and growth rates are normally low over winter. Increases with temperature in fish, up to optimum temps, e.g. for G trout Topt=13.1-14.1

oC (diet

dependent), for salmon Topt=15.9oC. Any growth potential resulting from metabolic rate increases would only be realised if invertebrate food is available.

Upper incipient critical limits for survival of juveniles are 22-33oC for salmon and 20-30 for brown trout.

Direct field observations are few and effects dependent upon food availability. The partial offset due to elevated early and reduced later temps may have accounted for Crisp’s observation that trout growth downstream of Cow Green reservoir were not affected when computed from 1

st Jan, using models. Same

procedure shows significant +ve effects on Wye salmon 0+ growth computed from early May (swim up) spring onwards due to net cooling effect at that site (Milner unpublished). Survival also temperature limited. E.g. trout upper limit 22-25

oC, salmon 22-28

oC. Modified by acclimation and genetics.

Crisp (1981, 2000) Elliott and Hurley (1989) Elliott and Elliot (2010) Jonsson and Jonsson (2009) Jensen (2003)


As above, G proportional to temperature up to Topt in all temperate fish species. Spring incubation times of coarse fish will be reduced by higher early spring temps and their growth will increase. Thermal tolerances are different between species, life stages, acclimation history, dissolved oxygen, pollutants and the location on the thermal scale. Salmonids generally are cold water species compared with cyprinids.

-


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See above. Egg incubation periods for salmonid eggs seem to cover both higher (winter) and lower (spring) temperature changes, approximately cancelling effects out (on the Wye, Craig Goch for example).Likely to be site-specific. Lethal or growth rate thresholds likely to be reached less often in zones subject to cooling. Delayed onset of bullhead growth, possibly due to lower spring temps in Tees downstream of Cow Green reservoir. At Kielder Water in Northumbria, a winter warming thermal effect in the absence of winter stratification caused adverse effects on salmonid egg hatching – too early due to warmer winter temperatures, followed by a delay in the spring temperature increase that will initiate feeding (particularly salmon as is 7

oC rather

than 4oC for trout), thereby causing starvation of alevins).

Crisp et al.(1983) Haile, James and Sear (1989)


Impacts of low dissolved oxygen concentration,, pH, metal sediments on fishes are well-established. Brooker reviewed these in relation to downstream effects of dams. Low dissolved oxygen concentration, could lead to problems for developing embryos ranging from death to delayed or abnormal development. Salmonids are more sensitive to low dissolved oxygen concentration, than cyprinids and eels. They are also more sensitive to low pH and the associated toxicity of aluminium. Acid water problems are still significant in upland catchments with poor buffering capacity, exacerbated in many upland cases by the compounding effect of coniferous forestry. H

+ toxicity is profoundly moderated by calcium and reservoir stored water has typically low base content (as might

be the case draining uplands (Brooker, 1981). Then the release of such water at times when river pH is low (e.g. pH<5.5) and the base cation concentration would normally be high (as in winter and spring), could lead to problems particularly for salmonid eggs and fry. No examples of this are reported, but it is a plausible mechanism to impact downstream fish populations. The reverse situation in which increased retention exacerbates the negative impacts of downstream acidic tributaries of of Co2 supersaturated groundwater has been reported to be a problem for salmon smolts in Scandinavia. No examples of significant reservoir related impacts on fishes due to impoundment effects on water quality have been reported in the UK.

Brooker (1981) Davies (1975) Rosseland and Kroglund (2011)


Limited information, but see above. Stored water undergoes changes including an increase in iron and manganese in hypoxic layers. This can lead to downstream precipitation and accumulation in sediments. The effects of invertebrates have been already noted; but the concretions that typically result could also render gravels unsuitable for the creation of redds and prevent the permeability necessary to successfully incubate fish embryos. There are no published examples of this effect in the UK.

-

High concentrations of algae and chlorophyll a, with low concentrations of dissolved nutrients (that have been utilized by the algae) and high pH.

Limited information, but see above. Increased nutrients in released water up to a point will enhance fish production; but if eutrophication proceeds too far, then low dissolved oxygen concentration, directly and through plant die off, can result in fish deaths. Conceivably the released water could have higher alkalinities than receiving water, and if in suitable ranges can lead to increase photosynthesis an cause the pH to rise above tolerance levels.

Brooker et al. (1978) Brooker (1981) MAFF (1972)


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3.5 What processes affect the influence of impoundments on downstream watercourses and their ecology?

This review has identified the four most significant risk factors for ecology in rivers downstream of large impoundments as a result of altered water temperature and chemistry. The majority of these risk factors are associated with hypolimnial releases from deep impoundments and should be carried forward towards the development of a protocol for mitigating the impacts of large impoundments on downstream river water bodies. These four risk factors are:

1. reduction in mean and / or maximum summer temperatures (summer depression); 2. increase in mean and /or minimum winter temperatures (winter elevation); 3. hypolimnial releases from stratified reservoirs with high concentrations of nutrients

and high concentrations of dissolved metals; and 4. releases from eutrophic reservoirs with high nutrient concentrations.

Webb (2008) emphasises that the expression of impoundment effects on rivers depends on processes occurring within reservoirs, in the downstream water courses and in the reservoir catchment. These processes include: climate; geography; size of impoundment; impoundment operation; and the character of downstream water courses (Webb and Walling, 1993a, 1996; Preece and Jones, 2002).This section of the report considers the processes that operate at each impoundment to determine how these four ecological risk factors manifest. It considers how impoundment characteristics can be used to form part of the protocol for mitigating the impacts of large impoundments on the ecology of downstream river water bodies.

Based on the findings of the literature review, and disregarding climatic effects responsible for inter-year variation, it is considered that the nature and scale of the effects of a given impoundment on downstream water chemistry and water temperature in temperate climates is largely influenced by the following processes:

physical characteristics including: altitude, retention time, exposure to wind, surface area to volume ratio, depth, land use and trophic status (broadly oligo-, meso- or eutrophic);

operational characteristics including: mode of operation, position, type and size of release(s) to downstream watercourse, including steady and variable release characteristics and spill characteristics; and

downstream watercourse characteristics including: water depth, slope, flow type, amount of groundwater baseflow (BFI), location and size of tributary inflows, riparian vegetation/tree cover, presence of other anthropogenic factors (water abstraction or discharges).

These processes are described in turn below.

3.5.1 Physical and operational impoundment characteristics

The presence of a natural lake or any impoundment will influence the downstream water temperature and chemical characteristics in comparison with the natural stream, even in the absence of stratification (for example, a winter warm effect is noted downstream of Kielder Water in the absence of stratification (Haile, James and Sear, 1989) and various effects are noted downstream of Wimbleball Lake in the absence of stratification (Webb, 1995).


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However, the most severe downstream effects that are most likely to have important ecological consequences are likely to occur as a consequence of hypolimnial releases from large and deep impoundments, and especially those that stratify. Both thermal and chemical effects tend to be exacerbated by reservoir characteristics that allow stable stratification to develop. As well as the physical characteristics of an impoundment, the characteristics of its operation may also affect the type and degree of effect it has on the downstream watercourse e.g. the position of the release.

Some ecologically damaging thermal and chemical effects may also arise in rivers from epilimnial releases or releases from small/shallow impoundments in eutrophic or polluted reservoirs.

Given that reservoir stratification has been identified as one of the major risk factors for adverse downstream thermal and chemical effects of impoundments, the causes and occurrence of stratification in UK reservoirs have been explored in more detail.

Gilvear et al. (2002) state that stratification of the water column is rare in Scottish impoundments and attribute this to high wind speeds. Although wind-induced destratification can be particularly effective for large surface area, shallow reservoirs (Petts, 1986), no references or data are cited to support the assertion of Gilvear et al. (2002) with respect to lack of stratification in Scottish impoundments. Based on the wider literature, it is considered that there is a risk of some kind of stratification in most deep impoundments in the UK, including those in Scotland. Within the temperate zone, thermal stratification during summer is common within reservoirs having a depth greater than 10 m (Petts, 1986), although the thermal behaviour of short-retention impoundments is largely controlled by the inflowing water mass (Arai, 1973). An additional winter stratification may be exhibited by northern latitude and high altitude reservoirs, where surface waters below 4oC lead to an inverse temperature gradient with maximum density conditions at the bottom (Petts, 1986). Furthermore, Ryan et al. (2001) investigated impoundments in the USA and suggested that any impounded water body greater than 5 m in depth can undergo thermal stratification, except those reservoirs where average yearly inflow volume exceeds the reservoir volume by a factor of 10 or more (Harlemann, 1982).

In the north of England, Crisp (1977) notes a striking difference between the effects of Cow Green reservoir on the downstream River Tees when compared with the effects of Selset and Grassholme reservoirs on the downstream River Lune; the summer maximum temperature was depressed by about 4oC in the River Lune but only 1-2oC in the River Tees below Cow Green reservoir. This difference could reflect differing timescales over which the data were collected (10-day means in the Lune, monthly means in the Tees) or differences in topography, aspect or other physical characteristics of the two sites (Crisp, 1977). Based on the storage capacity and yield of the reservoirs, it is likely that Cow Green has the longer retention time.

Together, Selset and Grassholme reservoirs have a smaller storage capacity than Cow Green reservoir, as well as a smaller combined surface area and higher maximum depths, suggesting that although their combined retention time is lower, they may be more prone to stratification. Lavis and Smith (1972) note that maximum stratification in the Lunedale reservoirs occurs during the summer months (May to August), and that the compensation release from Grassholme reservoir is drawn from 28.3 m below top water level (i.e. a deep water release).


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In contrast, water is drawn off simultaneously from the upper and lower levels of Cow Green reservoir (a mixed release), which, in combination with its relatively high altitude (489 m), high and frequent winds (average wind speed 24 km/h) and relatively shallow depth (max. 23 m) results in considerable mixing and infrequent stratification (Armitage, 2006). The pattern of flow release from Cow Green reservoir has remained similar since impoundment with maintained minimum flows in summer and occasional overtopping in winter (Armitage, 2006).

Crisp (1977) postulates that the presence of stratification at Lunedale reservoir and its relative scarcity at Cow Green reservoir could contribute to the observed difference in the depression of summer temperature maxima between the two sites.

Llyn Brianne is a deep (max. depth 85 m) regulating reservoir in the upper Tywi catchment in mid-Wales with a relatively small surface area (~22 ha) and a deep water discharge point (65 m below top water level). The reservoir is reportedly thermally stratified between May and October (Wyke, 1997), and this may be associated with its relatively small surface area to volume ratio and long retention time.

Kielder reservoir is usually well-oxygenated in winter due to wind-induced circulation, but stratification begins in early May and lasts until natural overturn in September and October (Haile, James and Sear, 1989). Even in the absence of stratification during the winter months, the authors report warmer downstream winter water temperatures, including adverse effects on salmonids.

At first glance, investigations of the River Haddeo (Webb and Walling 1998a, Webb 1995) suggest it is unlikely that thermal stratification of Wimbleball Lake and its hypolimnetic release has affected the thermal regime of the downstream river, because the reservoir is equipped with a de-stratification system and strong and persistent thermal gradients are generally absent. However, it does appear that construction of the dam substantially increased the flow of springs at its base (thought to be fed by water from the base of the lake). In the winter period the spring-flow contribution to the compensation flow is generally warmer than runoff from the reservoir or flow in the unregulated river and vice versa during the summer (Webb and Walling, 1996), thus the spring-flow appears to affect the downstream thermal regime in a similar way to that of a deep reservoir release (although the overall effect is the result of a complex interplay between increased spring-flow and reservoir releases).

The downstream effects of Wimbleball Lake are reportedly consistent with findings reported for other reservoirs in Britain that are relatively shallow and exhibit poorly developed stratification (Webb, 1995), and despite some of the observed biological effects being attributed to site-specific conditions it is considered that downstream impacts on invertebrate and trout development in the regulated river arise largely due to changes in water temperatures in midsummer, autumn and winter (Webb (1995). Nevertheless, it should be noted that the impact of Wimbleball Lake on downstream thermal regime is relatively modest compared with larger reservoirs and those in different climatic settings where temperature extremes have been altered to a much greater extent, annual temperature cycles have been more severely disrupted, and the effects of modification persist for much greater distances downstream (Webb, 1995).

Chew Valley Lake in Somerset (constructed 1953) has a surface area of 4.86 km2 at top water level, a mean depth of 4.27 m and a maximum depth of 11.5 m. Wilson et al. (1975a)


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showed that under natural conditions it would stratify intermittently during the summer, with anoxia rapidly developing in the hypolimnion and significantly reducing the quality of abstracted water. Artificial de-stratification was therefore introduced to maintain chemical water quality to the treatment works during the summer months (Hilton et al. 1992).

Soja and Wiejaczka (2014) report that thermal stratification occurs in the Klimkowka reservoir on the River Ropa in Poland during the summer months, causing releases 10-20oC cooler than upstream water temperatures due to a bottom water release. The reservoir is situated in the Carpathians, has a capacity of 43.5 million m3, a surface area of 3.1km2, is more than 5km long, and its maximum depth by the dam is 30 m. Downstream recovery of water temperatures is evident, but not to the extent observed in the reaches above the reservoir. In winter, the reservoir has a warming effect; the outflow temperature is 1.5-2oC higher than the inflow, and again, the effect is propagated downstream with reduced prevalence of icing (fewer days with ice cover following reservoir construction, although more days with partial or anchor ice cover). There is reduced magnitude of change between summer and winter temperatures i.e. amplitude of variation is dampened, on average by 5oC.

Reports by the Tay Foundation on the downstream thermal effects of impoundments on the River Lyon and River Errochty (Summers, undated a and b) indicate that spring and summer water temperatures immediately downstream of the dams displayed relatively little daily variation and were relatively cold, whilst in winter, temperatures were relatively warm. Some downstream amelioration was noted due to solar radiation and incoming tributaries. Interestingly, spring and summer water temperatures downstream of the dam at Lubreoch on the Lyon were observed to remain stable for periods and then to increase in a stepwise manner. The stable periods coincided with warm ambient air temperatures, whilst the step increases occurred when ambient temperatures appeared to reduce. This effect was therefore attributed to development of thermal stratification in Loch Lyon under prevailing warm and calm conditions, followed by mixing of the warmer epilimnion with the cooler hypolimnion during periods of cooler, windier weather, causing water temperatures at the bottom of the loch (from where water is released) to rise.

Stratification of Grafham Water is reportedly prevented for most of the year by its location, shallowness and high surface area, which enable wind action to maintain a well-mixed water column. Grafham Water is a eutrophic pumped storage reservoir, fed by abstraction from the nutrient rich River Great Ouse and discharging to Diddington Brook. Water level in the reservoir can fluctuate greatly and in the past this has led to a loss of marginal vegetation with consequent effects on water nutrient levels (NRA, 1993).

Rutland Water (the second largest reservoir in the UK by volume) is a eutrophic pumped storage reservoir fed by abstraction from the River Welland and with a deep release to the River Gwash. The reservoir is reportedly deep enough to stratify, but de-stratification systems are in place to ameliorate this. No downstream effects of water temperature modification have been reported in the literature. Some historic and localised impacts of iron-rich deposits have been reported in the River Gwash, immediately downstream of the impoundment. During the late 1980s and early 1990s, ferric sulphate was added to Rutland Water to reduce the concentration of blue-green algae. The precipitate formed a ‘ferric floc’ that deposited on the bottom of the reservoir and passed through the impoundment into the River Gwash, causing localised impacts on macro-invertebrates (Chris Extence, Environment Agency, pers. comm).


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Table 3.6 Impoundment characteristic risk factors for downstream water quality impacts (including occurrence of stratification)

Impoundment characteristic Risk factors for downstream water quality impacts Notes

Altitude The majority of UK reservoirs identified from the literature as having downstream thermal or water chemistry effects are situated at relatively high altitude (>200 m AOD) e.g. Wimbleball Lake, Llyn Brianne, Kielder Water, Cow Green reservoir, Lubreoch and Stroniuch reservoirs. This may simply be a consequence of the sites selected for investigation, and/or the fact that in the UK, the majority of impounding reservoirs tend to be situated in the upper areas of catchments at relatively high altitude. There are, however, several low altitude pumped storage reservoirs e.g. Grafham Water (45 m) and Rutland Water (85 m). Although downstream effects of these reservoirs are not widely reported, they are considered possible on the basis that Rutland Water has been known to stratify, and has a deep release to the River Gwash. Anecdotally, historical ferric sulphate dosing of Rutland Water has resulted in bottom reservoir sediments with high concentrations of metals, including iron. Ferric deposits have been noted in the River Gwash downstream, but are considered more likely to be associated with sediment transport and deposition during periodic scour valve releases (as opposed to continuous release of anoxic baseflow from the hypolimnion). Effects do not appear to persist downstream. It should be noted that the investigation of altitude effects at the lowland reservoirs described above is likely to be complicated by their long residence times, large surface areas and relatively low depths in the UK context.

Altitude data are readily available. Reservoirs at higher altitude may be associated with a higher risk of downstream effects. It is possible that any effects associated with altitude are actually driven predominantly by differences in air temperature and wind speed which affect stratification. However, it is considered that other characteristics are likely to be more important predictors than altitude.

Latitude/longitude, location. Downstream thermal effects at Wimbleball Lake in Devon (Webb, 1995), at Llyn Brianne in mid-Wales (Wightman et al. 1990, Wyke, 1997), at Kielder Water in Northumberland (Haile, James and Sear, 1989) and on the River Lyon in Scotland (Jackson et al. 2007) (amongst others) appear to indicate that latitude/longitude and/or location per se are not significant distinguishing factors in the UK – other factors are likely to be more important predictors.

Downstream effects are reported for reservoirs throughout the UK. This characteristic does not appear to be a useful predictor.

Maximum Depth NB maximum depth has been used here as it is more widely reported in the literature than average depth. Where depth data are unavailable, the height of impoundment can also be used as a surrogate for maximum depth.

Llyn Brianne (max. depth 85 m) thermally stratified between May and October, deep release with downstream temperature effects (Wightman et al. 1990, Wyke, 1997) and sedimentation effects resulting in elevated levels of manganese and iron in the sediments downstream (Wightman et al. 1990). Kielder Water (max. depth 52 m) is well-mixed in winter, stratified in summer. Winter warming effect even in absence of stratification (Haile, James and Sear, 1989), including adverse effects on salmonids. Wimbleball Lake is relatively shallow <50 m with a de-stratification system and strong and persistent thermal gradients are generally absent. Nevertheless, downstream thermal effects are reported (Webb and Walling 1998a, Webb 1995). It appears that construction of dam has substantially increased flow of springs at base of dam (thought to be fed by water from the base of the lake) (Webb and Walling, 1996). Rutland Water (max. depth 33 m) is reportedly deep enough to stratify (de-stratification systems are in place) and has a deep release to the River Gwash, but few documented downstream effects. Stratification recorded in Caban Coch reservoir in mid-Wales (max. depth 40 m, relatively shallow and exposed) (water temperature differences of 11

oC between surface and draw-off at 25 m depth) Brooker (1981), citing Thompson (1954), Hopper

(1978). Downstream water quality and biological effects have been reported. Stratification occurs in the Klimkowka reservoir (max. depth 30 m) in Poland with downstream thermal effects (Soja and Wiejaczka, 2014). Pontsticill Reservoir, south Wales, maximum depth 29.81m, stratification in summer – until October (APEM, 2015a) Crisp (1977) attributes the rare and short duration of stratification at Cow Green reservoir to the depth of the reservoir (max. depth 22.9 m) and its exposed position. However, even in the absence of stratification, downstream thermal effects were still reported. Llwyn-Onn reservoir, south-wales, maximum depth 22.28m, Stratification in summer – until October (APEM, 2015b) Stratification of Grafham Water (max. depth ~21 m) is reportedly prevented for most of the year by its location, shallowness and high surface area, which enable wind action to maintain a well-mixed water column. Intermittent stratification reported under natural conditions in Chew Valley Lake (Somerset) (max. depth 11.5 m, mean depth 4.27 m) (Hilton et al. 1992 citing Wilson et al. 1975).

Depth data are readily available. In the UK literature, downstream ecological effects are documented for impoundments 22.9 m and deeper. The effects of impoundment depth on downstream water quality are likely to represent a continuum, with the deepest impoundments presenting the greatest risk of significant downstream ecological effects. The risk will be increased for impoundments where stratification occurs. The literature and APEM experience indicates that the deeper the reservoir, the greater the risk of stratification. Ryan et al. (2001) suggest that there is a risk of stratification for impoundments greater than 5 m, whilst in the UK context, Petts (1986) suggests that impoundments with depth greater than 10 m are at risk of stratification. In a screening exercise to identify potential effects downstream of Australian impoundments, Preece (2004) ruled out any impoundments shallower than 10 m on the basis that they were less likely to maintain stable and persistent stratification through warmer months in temperate climates. Whilst recognising that it represented a simplification of a complex process, Preece (2004) used structure height (as a proxy for depth and potential for stratification) as a key indicator in his assessment. Preece (2004) characterised the potential severity of downstream thermal (summer cooling) disturbance from impoundments as


March 2015 Page 46


Mackie et al. (1983) noted that changing from a surface to deep release had little or no impact on thermal regime downstream of the dam at Guelph Lake USA (max. depth 12.2 m, 400 ha surface area). Anecdotal information suggests that reservoir stratification and associated downstream effects are minor in Yorkshire, UK, where the majority of reservoirs are typically 10-40m deep with relatively deep outflows. Such reservoirs are probably also quite exposed to high wind speeds, and situated at relatively high altitude. Within the temperate zone, depths >10 m associated with summer stratification Petts (1986). Manchester Reservoirs UK: stratification occurring at the deeper reservoirs where depth >10m (APEM, pers. comm.). Pentwyn reservoir, south Wales, maximum depth 7.48, no stratification (APEM, 2015a) USA water bodies >5 m can undergo stratification. Ryan et al. (2001).

follows:

Severe deep intake (≥10 m) and large discharge (≥1000 Ml/d).

Moderate deep intake (≥10 m) and smaller discharge (>5 and<1000 Ml/d).

Minor shallower intake (>5 m and <10 m) and smaller discharge (>5 and <1000 Ml/d).

Negligible very shallow intake (≤5 m) and very small discharge (≤5 Ml/d).

Discharge was included on the basis that thermal characteristics of hypolimnetic releases will extend further downstream under high discharge because larger volumes of water are less responsive to heat flux and move at higher velocity (Preece, 2004; Ward, 1985). This characteristic is likely to be a useful predictor of effects, especially when used in conjunction with others.

Surface area In the UK context, Llyn Brianne and Selset reservoir have relatively small surface areas (22 ha and 111 ha, respectively, the former being particularly small) but both have documented downstream thermal and ecological effects (Wightman et al. 1990, Wyke, 1997), Lavis and Smith (1972)). In contrast Kielder Water has one of the largest surface areas of reservoirs in the UK (1086 ha) and also has documented downstream effects despite an apparent lack of stratification (Haile, James and Sear, 1989). Stratification of Grafham Water is reportedly prevented for most of the year by its location, shallowness and high surface area (628 ha), which enable wind action to maintain a well-mixed water column (NRA, 1993). Downstream effects are not reported. Data from recent APEM surveys (e.g. APEM, 2015ab) do not indicate a relationship between surface area per se and stratification.

Surface area data are readily available. In the UK literature, downstream ecological effects are documented for reservoirs with both very small and very large surface areas (e.g. Llyn Brianne, Kielder Water). Reservoirs with large surface area may be more prone to wind-induced mixing, but this will also depend to a large degree on reservoir volume and depth. APEM experience indicates the lower the surface area to volume ratio, the greater the risk of stratification. This characteristic may be a useful predictor of effects but only when used in conjunction with others.

Volume In the UK context, some of the largest biological effects appear to occur downstream of reservoirs with high volumes (e.g. Llyn Brianne, Kielder Water, Loch Lyon). However, no downstream effects are reported for Grafham Water or Rutland Water, both of which also have high volumes. Webb et al. (2008) cite a study in the former Czechoslovakia (Patera and Votruba, 1996) which suggests that relative reservoir capacity, expressed as the time taken for a reservoir to fill, was a useful discriminator of the effects of an impoundment on river temperatures. Data from recent APEM surveys (e.g. APEM, 2015ab) indicate stratification typically occurring at sites with the greatest volume. Kielder Water is the largest reservoir by volume in the UK (199 Mm

3), and despite a lack of reported stratification, downstream

effects have been reported (Haile, James and Sear, 1989). Rutland Water is the second largest reservoir in the UK by volume (124 Mm

3) and can stratify (de-stratification systems are in

place). However, downstream effects are not widely reported, except for very localised flocculated ferric deposits (Chris Extence, Environment Agency, pers. comm). Effects are reported downstream of Wimbleball Lake (21.5 Mm

3) and the Lunedale reservoirs (Selset (15.3 Mm

3) and Grassholme

reservoir (6 Mm3), despite their relatively small volumes.

Volume data are readily available. In the UK literature, downstream ecological effects are documented for impoundments of volume ~20 Mm

3 and greater.

Larger reservoirs may be more prone to stratification, but this will also be affected by depth and surface area, among other factors. APEM experience indicates that the lower the surface area to volume ratio, the greater the risk of stratification. This characteristic may be a useful differentiator of effects but only when used in conjunction with others.

Surface area to volume or surface In the UK context, environmental effects appear to occur downstream of reservoirs with low surface area to volume ratio (<10) (e.g. Surface area to volume and surface area to depth ratios can be


March 2015 Page 47


area to depth ratio Llyn Brianne (0.34), Kielder Water (5.46)). Of the UK reservoirs for which data were identified, Grassholme reservoirs and Loch Lyon are those with the highest surface area to volume ratios (9.41 and 8.34, respectively) where downstream effects were also reported. Fewer / less severe effects are reported downstream of reservoirs with high surface area to volume ratio (>10) (e.g. Chew Valley Lake, Rutland Water, Grafham Water); however, it is important to note that such lack of effect may be wholly or partially due to efficient artificial de-stratification systems at these sites. Data from recent APEM surveys in Wales and north west England indicate stratification typically occurring at sites with the lowest surface area to volume ratio. Manchester Reservoirs (UK) data indicate stratification occurring at sites with lower surface area to volume ratio (APEM pers. comm.). These were however also the sites with the greatest water depths. Delayed and reduced summer peak temperature, and slightly elevated temperatures during the winter are reported downstream of Llyn Brianne which has a low surface area to volume ratio (22 ha/64.4 Mm

3, 0.34) (Wyke, 1997).

Effects are also reported downstream of Kielder Water (1086 ha/199 Mm

3, 5.5) and Caban Coch reservoir (202.3 ha/35.5 Mm

3,

5.7) which have relatively low surface area to volume ratios in the UK context. Pontsticill reservoir, south Wales (104 ha/13 Mm

3, 7.8) stratification in summer until October (APEM 2015a). No information

available on downstream effects. Llyn-Onn reservoir, south Wales also has a relatively high surface area to volume ratio (55 ha/5 Mm

3, 10.6) and stratification

reported in summer until October (APEM, 2015b). No information available on downstream effects. Intermittent stratification was reported in Chew Valley Lake (Somerset) under natural conditions despite a relatively shallow depth and very high surface area to volume ratio (486 ha/20 Mm

3, 23.7).

readily calculated from available surface area, volume and depth data. APEM experience indicates the lower the surface area to volume ratio, the greater the risk of stratification. Downstream effects may be greater downstream of reservoirs with lower surface area to volume ratio but other factors are also likely to be important. This characteristic may be a useful predictor of effects when used in conjunction with others. Downstream effects were reported for reservoirs with both very small and very large surface area to depth ratios (Llyn Brianne (0.26), Kielder Water (20.88)) in the UK context. This characteristic appears unlikely to be a useful predictor.

Retention time (incl. size of reservoir, inflow and outflow)

Hypolimnial releases: Only limited data were available on retention time in UK reservoirs within the literature. The Lunedale reservoirs: Selset and Grassholme, have a relatively short retention time ~120 days, whilst Cow Green reservoir has a longer retention time ~264 days. Mean summer temperature depression was actually greater downstream of the Lunedale reservoirs than downstream of Cow Green reservoir; however, this could be due to differences in the occurrence of stratification at the two sites and average wind speed. There is also a waterfall immediately downstream of Cow Green reservoir. Grafham Water has a long retention time 248.2 days and Rutland Water has a very long retention time 766.5 days (Wilson, 1995). Neither have reported downstream effects due to water temperature modification; only localised impacts of flocculated iron deposits caused by historic dosing of the reservoirs with ferric sulphate (Chris Extence, Environment Agency, pers. comm). Reservoirs where average yearly inflow volume exceeds the reservoir volume by a factor of 10 or more are reportedly not prone to stratification (Harlemann 1982). The thermal behaviour of short-retention impoundments is largely controlled by the inflowing water mass (Arai, 1973). The extent of deterioration in water quality associated with stratification will be related to the retention time of the reservoir, with longer storage times leading to increased deterioration. This is usually attributed to the effects of poor mixing (Grabowska, 2002). Soares et al. (2008) attribute stable stratification in a Brazillian reservoir to long retention time and low wind speed (<5 m/s).

Retention time data are not readily available. Reservoirs with long retention times may be more prone to downstream thermal and chemical effects and to stratification, however, other factors are also likely to be important. This characteristic may be a useful predictor of effects but only when used in conjunction with others.

Epilimnial releases: Conceptually, there may be greater risk of effects associated with surface releases from small impoundments where such impoundments have long retention times, which may facilitate summer warming. Lessard and Hayes (2003) indicate that d/s effects of surface releases from small dams (<5 m) are affected by surface area, depth and retention time, but thresholds for these are not provided. Stable stratification may also facilitate summer warming of the epilimnion. However, little documented evidence was identified to

Retention time data are not readily available. Reservoirs with long retention times may be more prone to downstream thermal and chemical effects and to stratification, however, other factors are also likely to be important. This characteristic may be a useful predictor of effects but only when used in conjunction with others.


March 2015 Page 48


support this. Longer retention time promotes phytoplankton growth (Grabowska, 2002).Grabowska (2012) reports that a shallow eutrophic reservoir (Siemianówka Dam Reservoir (sic.) – max. water level 145 m above sea level, volume 79 Mm

3, surface area 3250 ha,

altitude 147 m above sea level) leads to deterioration of water quality in the lowland river basin through changes in the phytoplankton composition and abundance. This was due to dominance of cyanobacteria. Water retention time, daily water outflow and TN:TP ratio were identified as the factors favouring cyanobacteria development in the reservoir with conditions in the river downstream also favouring development (meandering character and low flow velocities in the outflowing river, together with the method of operating the dam gates). The study demonstrated that water retention time exceeding three months in shallow dam reservoirs significantly increases the risk of river water quality deteriorating as a result of mass development of cyanobacteria.

Exposure to wind, orientation to prevailing wind direction and topography (aspect, degree of sheltering and shading etc.)

Only limited data were available on wind-induced mixing effects in UK reservoirs within the literature. Wind-induced de-stratification can be particularly effective for large surface area, shallow reservoirs Petts (1986) Gilvear et al. (2002) attribute the apparent lack of stratification in Scottish reservoirs to high wind speeds.

Crisp (1977) attributes the rare and short duration of stratification at Cow Green reservoir to the depth of the reservoir (22.9 m) and

its exposed position. Stratification has been recorded in Caban Coch reservoir in mid-Wales, despite its relatively shallow depth (max. depth 40 m) and

exposed position (Brooker, 1981). Soares et al. (2008) attribute stable stratification in a Brazillian reservoir to long retention time and low wind speed (<5 m/s).

Wind speed data may be readily available at some sites. Reservoirs with lower exposure to wind (due to any of altitude, orientation, topography, land use, surface area etc.) may be more prone to downstream effects and to stratification, however, other factors are also likely to be important. This characteristic may be a useful predictor of effects but only when used in conjunction with others.

Trophic status (broadly oligo-, meso- or eutrophic) and land use

Hypolimnial releases: Downstream effects are reported for UK reservoirs with both oligotrophic and eutrophic status. Conceptually, downstream effects may be exacerbated for eutrophic impoundments, as these would be expected to have bottom sediments with higher nutrient content due to settled material. However, little documented evidence was identified to support this.

Information on trophic status may be readily available at some sites. Downstream effects may be exacerbated below eutrophic reservoirs, but this is considered uncertain. This characteristic may be a useful predictor of effects but only when used in conjunction with others.

Epilimnial releases: Conceptually, there may be greater risk of effects associated with surface releases from small impoundments where the reservoir is eutrophic and more algal development is possible. Grabowska (2012) reports that a shallow eutrophic reservoir (Siemianówka Dam Reservoir (sic.) – max. water level 145 m above sea level, volume 79 Mm

3, surface area 3250 ha, altitude 147 m above sea level) leads to deterioration of water quality in the

lowland river basin through changes in the phytoplankton composition and abundance. This was due to dominance of cyanobacteria

Information on trophic status may be readily available at some sites. Downstream effects may be exacerbated below eutrophic reservoirs. This characteristic may be a useful predictor of effects for small, surface release reservoirs when used in conjunction with others.

Use/operation

Conceptually, pumped storage reservoirs may have more variable water levels and/or longer retention times c.f. Grafham and Rutland Water. However, little documented evidence was identified to suggest downstream effects as a consequence of this.

This characteristic does not appear to be a useful predictor.

Position of release

In Ontario, Mackie et al. (1983) noted that changing from a low to high discharge level had little or no impact on thermal regime downstream of the dam. A fixed surface or deep release is a potential risk for downstream water quality effects. Fixed deep releases from large impoundments are likely to be particularly problematic. The presence of a variable or surface release is likely to reduce the potential for and severity of downstream effects. A surface release may be less operationally costly than a variable release mechanism. The more flexible the release mechanism and the better the operational management, the less likely downstream effects are to occur. However, it is likely that even very flexible releases will not be able to completely remove all downstream effects in certain

This characteristic is likely to be a useful predictor of risk of downstream effects.


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circumstances. Such mechanisms may also be very costly (capital and operational). It is possible that the presence of a hypolimnetic draw-off to supply or to the downstream watercourse may induce mixing within the reservoir, and therefore confer a benefit to in-reservoir water quality, for example, by reducing the risk of stratification. In some circumstances, changing from a deep to a surface release may therefore pose a risk of deterioration of in-reservoir water quality.

Type of release regime Downstream temperature and water chemistry are intimately linked to the downstream discharge regime. A higher volume hypolimnial baseflow is likely to have greater downstream effects than a lower volume hypolimnial baseflow, such that regulating releases, hydro-peaking releases and variable base-flows may be more problematic than a low volume base-flow if taken from the hypolimnion. Hydro-peaking releases may cause rapid changes in downstream water quality and are therefore likely to be undesirable. Ramping of such releases may confer some mitigation. Flow from unregulated catchment areas may ameliorate some effects of hypolimnial baseflow releases. The effects of the spill regime (surface release) will depend on the chemical characteristics of the reservoir and the frequency, magnitude and duration of spills. Re-wetting a reach with hypolimnial base-flow may introduce undesirable water quality effects, particularly in the reach immediately below the dam, before introduction of tributary inflows. If, for example, manganese deposition is induced immediately downstream of the dam, infrequent spills may periodically move deposits to reaches further downstream. At Cow Green reservoir, the compensation flow increases minimum flows above those naturally expected, and maximum flows are considerably reduced/dampened, even when uncontrolled flow is occurring over the spillway. (Crisp, 1977) Lowney (2000) noted that having a varying rather than stable release prevented the formation of downstream nodes and antinodes of minimum and maximum diurnal variation. Ellery and Wilkins (1999) report that mean temperatures in the River Elan immediately downstream of Caban Coch reservoir were similar during a regulation release (higher flow 2.677 m

3/s) and a compensation release (lower flow 0.812 m

3/s), although diurnal

variation appeared slightly lower under the higher flow. Under both release regimes a demonstrable decrease in mean temperatures in the River Wye was observed approximately 8km d/s of the dam. At the higher flow, less warming occurred along the length of the River Elan between the release and the confluence with the River Wye, and the regulation release exerted a greater influence on temperatures in the Wye than did the lower compensation flow. The downstream point at which the regulation release had an influence on temperature was approximately 20 km d/s of the dam. Lower DO concentrations were recorded during the regulation release immediately below the dam, but increased with distance downstream and the release did not affect concentrations in the River Wye under compensation or regulation conditions. The average pH immediately below the dam was approximately 1.2 units lower than that measured on the River Wye upstream of the confluence during both compensation and regulation. Along the River Elan between the dam and the Wye confluence, pH increased by 1.30 units on average during both compensation and regulation. Wightman et al. (1990) note that under normal summer conditions, reduced river temperatures downstream of Llyn Brianne extended as far as Llandovery (~20 km downstream), although extension much further downstream may occur during larger scale releases. A river management release in the summer of 1983 reduced river temperatures from 20

oC to 13

oC at Manorafon, 36 km

downstream of the reservoir. A recent survey in high summer, indicated reduced water temperatures for up to 25 km downstream

of Llyn Brianne (APEM, 2013). Gilvear et al. (2002) note that scour valve releases and other types of reservoir releases (e.g. freshets) can be important in affecting river water quality downstream, due to differences in the chemical composition of the reservoir water and the mobilisation of in-channel stores. Foulger (1986) monitored chemical changes during a 15 m

3/s release on the River Garry draining to Loch

Ness and found that, although chemical dilution was the principal water quality response, a small increase in the concentration of calcium was observed on the rising limb of the flow increase (Gilvear et al. 2002).

Significant effects arising as an indirect consequence of the quantity and dynamics of the flow release regime have been disregarded, since it is assumed that if flow regime mitigation measures are in place (i.e. an appropriate release regime which incorporates the required flow components as feasible/cost effective at a given site) then downstream heating and cooling effects should approximate those of the natural watercourse.

Reducing overly high hypolimnial baseflow releases is likely to be of particular importance for reducing downstream summer cooling effects.


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Size of release compared to receiving watercourse

For gradients in downstream effects to be established, the reservoir release must be of sufficient magnitude to overcome the influence of tributary inputs and ambient heating within the channel (Petts, 1986). Downstream effects will be greatest in re-wetting scenarios and reaches with minimal unregulated tributary inputs. Small releases to large receiving watercourses are likely to have the lowest effects, except in the situation of an anoxic hypolimnial release to a sluggish downstream watercourse.

Preece (2004) characterised the potential severity of downstream thermal (summer cooling) disturbance from impoundments as follows:

Severe deep intake (≥10 m) and large discharge (≥1000 Ml/d)

Moderate deep intake (≥10 m) and smaller discharge (>5 and<1000 Ml/d)

Minor shallower intake (>5 m and <10 m) and smaller discharge (>5 and <1000 Ml/d)

Negligible very shallow intake (≤5 m) and very small discharge (≤5 Ml/d)

Discharge was included on the basis that thermal characteristics of hypolimnetic releases will extend further downstream under high discharge because larger volumes of water are less responsive to heat flux and move at higher velocity (Preece, 2004; Ward, 1985). This characteristic is likely to be a useful predictor of effects, especially when used in conjunction with others.

Presence of artificial de-stratification systems

Conceptually, presence of artificial de-stratification mechanisms is likely to reduce downstream impacts. This may explain the lack of documented effects downstream of Rutland Water.

This characteristic is likely to be a useful predictor of risk of downstream effects, especially when used in conjunction with others.


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3.5.2 Downstream watercourse characteristics

It is important to recognise that downstream effects will also be influenced by the nature of the downstream catchment, for example, the size and position of downstream tributary inputs, and the character of the receiving watercourse (Table 3.7). Although downstream watercourse characteristics are considered below, they are not explicitly incorporated into the decision framework since it is targeted at prioritising high risk impoundments.

Table 3.7 Watercourse characteristic risk factors for downstream water quality impacts

Watercourse characteristic

Risk factors for downstream water quality impacts

Location and size of tributary inflows

Conceptually, low volume and/or distant tributary inputs are likely to mean that any release from an impoundment will have a larger effect on the downstream watercourse.

Amount of groundwater baseflow (BFI), amount of hyphoreic and phreatic exchange

Conceptually, low BFI or low volume of groundwater baseflow inputs are likely to mean that any release from an impoundment will have a larger effect on the downstream watercourse.

Slope, flow type, substrate and morphology

A low energy downstream watercourse is likely to reduce the rate of re-aeration and equilibration with air temperature. A high energy downstream watercourse is likely to increase the rate of re-aeration and equilibration with air temperature, although this will not resolve other issues such as high or low nutrient concentrations. A morphologically varied downstream watercourse is likely to provide refuges for biota, such that the effects of problematic releases on ecological status may be lower. Conversely, a morphologically uniform reach provides fewer refuges such that the effects of problematic releases on ecological status may be greater (Dunbar et al. 2010). Over-deepened and narrowed channels may take longer to equilibrate with air temperature, particularly for ‘summer cool’ releases. Braided channels may shed heat less easily, particularly for ‘summer warm’ releases.

Water depth and wetted width

A low depth to wetted width ratio in the downstream watercourse may facilitate downstream equilibration with air temperature and /or warming of cold releases in the summer months. Conversely a low depth to wetted width ratio in the downstream watercourse may exacerbate ‘summer warm’ effects if excess heat load cannot be shed.

Riparian vegetation/tree cover (shading), and/or topographical shading

Shading may extend the time taken for warming of cold releases in the summer months. Shading may facilitate shedding of excess heat from ‘summer warm’ releases during the summer months.

Presence of other anthropogenic factors (e.g. water abstraction or discharges, thermal effluent)

Water abstraction may ameliorate ‘summer cool’ effects as a lower volume of water will equilibrate more easily with air temperature, but may exacerbate ‘winter warm’ and ‘summer warm’ effects. Thermal discharges may ameliorate ‘summer cool’ effects to some extent, but may exacerbate ‘winter warm’ and ‘summer warm’ effects.


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Of the factors that will affect downstream thermal equilibration (or lack of), Caissie (2006) indicates that heat exchange occurs mainly at the air/water interface (82%) vs the streambed/water interface (15%) and via other processes (6%) (although this distribution is reportedly less clear for smaller streams). When documenting such effects it is important to note that measurement scale is also important; although water temperature is highly influenced by air temperature, different time scales will yield different air to water temperature relationships (Caissie, 2006).

Water depth and flow velocity influence the rate at which stream temperature will come into equilibrium with changed atmospheric conditions in the downstream direction (Webb et al. 2008). A study of small streams in the southern Swiss Alps (Meier et al. 2003) indicated that water diversion for HEP generation would only have a minor effect on the temperature of very steep river sections where dissipation of kinetic heat energy is the dominant process, but would be more significant for reaches of lower slope (Webb et al. 2008). Lack of downstream habitat variability may limit the availability of refuges from temperature and water quality effects.

In this context it is interesting to note that although the presence of Cow Green reservoir has altered the downstream temperature regime of the River Tees, downstream concentrations of dissolved oxygen were not appreciably affected (Crisp, 1977). This may be associated with the aerating effect of the Cauldron Snout waterfall downstream of the dam; shortly downstream of Cow Green, the River Tees drops about 40 m in 135 m (Armitage, 2006).

In the River Haddeo downstream of Wimbleball Lake, downstream temperature recovery was found to be associated predominantly with heat exchange with the atmosphere, and to a lesser extent with tributary inflows (Webb, 1995), although changes in temperature associated with tributary inflows also showed clear diurnal variation. Because the temperature changes in the tributaries tended to be out of phase with those in the main stream, tributary inflows acted to cool water parcels moving downstream during the day but warm them at night. In consequence, the effect of atmospheric heat exchange on downstream recovery was somewhat offset by those of tributary inputs (Webb, 1995).

The size and location of tributary inputs is also an important consideration; the impact of the deep water discharge from Llyn Brianne on river temperature is noted to be dampened with distance downstream of the dam, because the water re-equilibrates with air temperature and is diluted by the inflowing of unregulated tributaries (Wyke, 1997). Similarly, Krause et al. (2005) note that downstream of Philpott Dam in southwestern Virginia (USA), water temperatures in the river downstream exhibit little diel variation close to the dam, and that the extent of downstream effects on water temperature would be greater during summer (June-August) if not for inflows from a tributary at 5.3 km downstream of the dam. Poole and Berman (2001) provide a matrix of the relative influence of stream characteristics on temperature in small, medium and large streams (Figure 3.4).


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Figure 3.4 Relative influence of stream characteristics on temperature in small, medium and large streams from Poole and Berman (2001)

Webb et al. (2008) also comment that consequences of agricultural practices (in particular livestock grazing) for thermal regime has been the subject of recent studies inferring significant elevation of stream temperature with disturbance of the riparian environment through vegetation removal and stream widening and shallowing (citing Li et al. 1994, Quinn et al. 1997, Belsky et al. 1999, Isaak and Hubert, 2001). Despite this Webb et al.(2008) acknowledge that the influence of land use has been considered small compared to that of weather conditions (citing Borman and Larson 2003).


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4. Risk assessment decision framework

A key objective of this report is to advise on the development of a practical tool that can help regulators and water managers assess the risk of adverse impacts on water quality (temperature and chemistry) and ecology downstream of impoundments. This section of the report synthesises the information presented in the literature review into a decision framework to help identify where there might be risks of adverse ecological effects from water temperature and chemistry effects at UK impoundments and how these effects might be mitigated.

It should be noted that this decision framework is necessarily a simplification of very complex processes, nevertheless it is a risk-based approach which can be used to categorise impoundments in terms of likely magnitude of downstream impact, and sets out the questions that need to be asked at a given site. Where detailed information regarding stratification and downstream ecological impacts are available for a particular site, it is recommended that the UK regulators use such information to override the decision framework presented below, to arrive at a site-specific solution. An adaptive management approach is recommended following implementation of any mitigation measures (Allen et al. 2011; Rist et al. 2013).

The biggest risks are associated with hypolimnial releases from large, deep impoundments, particularly those that stratify (>5-10 m max. depth2). Of the effects of hypolimnial releases, ‘summer cooling’ effects are likely to be most problematic, and particularly for invertebrate populations. Such effects may be of more significance for coarse fish populations downstream of lowland reservoirs than for salmonid populations downstream of upland reservoirs. It is assumed that resolving the problems associated with the temperature effects of hypolimnial releases will also resolve any water quality impacts (i.e. temperature effects have been used as a proxy, since significant water chemistry effects are driven by the same processes). In such cases, it cannot be assumed that modifications to the flow regime in line with UKTAG (2013) will be sufficient to negate any downstream water temperature and chemistry effects.

Based on the literature reviewed, impacts associated with hypolimnial releases from large, deep reservoirs are likely to be relatively short-lived. Effects are likely to be significantly attenuated by the addition of the first major downstream tributary, and in the UK, the majority of studies indicate a return to pre-impoundment conditions within ~20-30 km. In most situations, impacts are expected to be restricted to the first (and potentially second) WFD water bodies downstream of the impoundment.

Some ecologically damaging thermal and chemical effects may also arise as a consequence of releases from small/shallow impoundments in eutrophic or polluted reservoirs.

The effects of releases from eutrophic reservoirs may persist for longer distances downstream, but the literature relating to such impacts is relatively sparse.

2 NB maximum depth has been used here as it is more widely reported in the literature than average

depth. Where depth data are unavailable, the height of impoundment can also be used as a surrogate for maximum depth.


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Water chemistry and temperature effects of surface releases (and knock-on ecological effects) have not been considered in this risk assessment decision framework based on the literature review findings. There are two reasons for this: first, the effect of surface releases on downstream river water bodies resembles the thermal and chemical characteristics of natural lakes; and second, there is a paucity of reports of ecological impacts from surface water releases compared to deep water releases.

On the basis of this literature review, a simplified decision framework has been developed to aid the UK regulators in assessing the potential water quality risks of releases from impoundments. This is presented in Figure 4.1 and utilises proxy variables as indicators of potentially significant effects. The proxy variables have been selected based on the literature review evidence, the value of the variables as indicators and on likely data availability. Other factors could be included to improve the weight of evidence at a specific site (e.g. low surface area to depth ratio, long retention time, high volume and low exposure to wind). However, data relating to these characteristics are not always readily available, and therefore they have been excluded from the main framework.


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Figure 4.1 Risk assessment decision framework

YesIs the reservoir

eutrophic?

No

Is the reservoir<5m max. depth?

Yes

No

Risk of downstream water chemistry and ecology effects associated with high nutrient concentrations. Alongside other measures, consider catchment management measures where eutrophication is a consequence of anthropogenic activity.

Very low risk of stratification and therefore very low risk of downstream water quality or ecology effects, regardless of position of release. Only investigate further if site specific information suggests this is essential. Adaptive management following implementation of mitigation measures.

Is the reservoir<10m max. depth?

No

Yes

Low risk of stratification and therefore low risk of downstream water quality or ecology effects. Only investigate further if site specific information suggests stratification occurs frequently and/or other site specific info indicates potential risks. Adaptive management following implementation of mitigation measures.

Risk of stratification

Is the release point:

Fixed

VariableAdaptive management following implementation of mitigation measures.

Is the release point:

Deep

SurfaceAdaptive management following implementation of mitigation measures.

Risk of downstream water quality effects

that might affect ecology

Is it possible to implement alternative options for delivery of

downstream flow?

No

Yes

Consider alternative options for delivery of downstream flow e.g. surface release, tributary diversion, variable release.

Consider alternative mitigation measuresEnsure flow regime is no higher than would naturally be expected during the summer monthsEnsure receiving watercourse is morphologically diverseConsider artificial resevoir destratification.


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In arriving at the simplified framework presented above, the following assumptions have been made:

The term ‘eutrophic’ in this instance is intended to be applied in a qualitative manner based on local knowledge of a particular site. It is included because eutrophic status may exacerbate the effects of impoundment on downstream water quality. Where eutrophication is identified, it will be important to understand if this is due to natural or anthropogenic factors, since in the former case, catchment management measures may be not be necessary, effective or appropriate.

The complexities associated with climate change and climatic effects responsible for inter-year variation have been disregarded.

Significant effects associated with the sediment transport regime have been disregarded since it is assumed that these will be dealt with separately i.e. via sediment management plans and modifications to the flow release regime.

Significant effects arising as an indirect consequence of the quantity and dynamics of the flow release regime have been disregarded, since it is assumed that if flow regime mitigation measures are in place (i.e. an appropriate release regime which incorporates the required flow components as feasible/cost effective at a given site) then downstream heating and cooling effects should approximate those of the natural watercourse (reducing overly high hypolimnial base-flow releases is likely to be of particular importance for reducing downstream summer cooling effects).

It is assumed that any temperature issues associated with hydro-peaking releases (i.e. rapid and untimely fluctuations in water temperature) would be resolved by modifications to the flow regime itself (as opposed to mitigation measures relating to the temperature of the flow regime). NB hydro-peaking releases are not thought to be a major issue in the UK (Michael Wann, SEPA, pers. comm.).

4.1 Mitigation Measures

A number of potential mitigation measures were identified during the global literature search and these are briefly summarised below (after Sherman, 2000; Olden and Naiman, 2010; and Petts, 1986):

1) Exploit reservoir stratification by selective withdrawal of water with the desired water quality

a. Surface releases b. Multi-level release structures c. Trunnions (floating intakes)

2) Submerged weirs or curtains (to provide a barrier to the passage of water and force warm or cold water above or below the curtain, respectively. Modifying the topography of the channel feeding into the dam can also result in a similar effect.

3) Stilling basins used to delay the downstream release of water so that thermal equilibrium may be reached with the atmosphere.

4) Artificial destratification to maintain isothermal conditions throughout the water column and prevent development of anoxic conditions at depth

a. Petts (1986) notes that air diffusion systems and mechanical pumping procedures have been successfully applied to small and medium sized lakes.

Olden and Naiman (2010) suggest that selective withdrawal using a multi-level intake structure is the most common and effective means of controlling the water temperature of dam releases. For example, Flaming Gorge Dam (Green River, USA) was installed with a


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multi-level intake structure in 1978 with the goal of increasing summer water temperatures, and partial success of this mitigation is evident in comparison of data collected before and after the measures were implemented (Olden and Naiman, 2010). The authors also note that reasonable improvements were observed in relation to the frequency and duration of thermal events, and the timing of extreme temperatures. In another example, Shasta Dam (Sacramento River, USA) was retrofitted with a multi-level intake structure in 1997 to improve downstream temperatures (Olden and Naiman, 2010). Dam management at this site is apparently focussed on releasing warmer surface waters in the winter/spring and cold deep waters in the summer/autumn, although the authors do not comment on the level of success achieved by these measures. The authors note that although most selective withdrawal intake structures are built during initial reservoir construction, release structures can also be successfully modified for selective withdrawal later following initial construction, subject to cost-benefit analysis (Sherman (2000) noted that retrofitting Australian dams with multi-level outlet structures was found to be unacceptably expensive for the benefits it accorded).

Surface releases are likely to be the preferred option for re-wetting scenarios, particularly if they are less costly and complex to operate than selective withdrawals. The literature indicates that a switch from deep to surface release of water from impoundments is likely to have beneficial effects on downstream water chemistry, temperature and ecology, though the scale of effects may vary widely on a case-by-case basis. For example, Mackie et al. (1983) noted that changing from a low to high discharge level had little or no impact on thermal regime downstream of the dam. Mackie et al. (1983) citing Stroud and Martin (1973) note that withdrawal of water from the top of the dam instead of from the bottom is usually at least partly responsible for effecting the development of an anoxic zone, as well as altering the thermal profile.

Therefore, site-specific investigation of the potential effects of changing the release depth may be advisable for large, deep impoundments with an existing hypolimnial release (particularly where a change to a surface release is being considered). This would include balancing potential downstream water quality and ecological quality benefits against potential exacerbation of in-reservoir stratification effects.


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5. Evaluation of Water Framework Directive environmental standards

UKTAG (2008b) identified that release of water to rivers from the cold depths of reservoirs might result in reduced downstream river temperatures and adverse effects on ecology and proposed that a maximum allowable temperature drop be applied that mirrored the maximum uplift values for heated thermal discharges to rivers. A maximum allowable reduction of 3°C at the edge of the mixing zone was therefore recommended for all cases except for waters of high ecological status. In the latter case a maximum allowable drop of 2°C was proposed. These standards relate to the ambient river temperature as an annual 98 percentile.

The findings of the literature review indicate that existing WFD environmental standards are likely to remain appropriate in the context of cold water releases from impoundments. The literature suggests that a 2-3oC cooling and warming effect of mean water temperatures during the summer and winter months, respectively, may have important negative effects on ecology (particularly for macro-invertebrates). Such changes are already broadly covered by existing standards. Winter warming effects are not considered likely to be ecologically significant in the UK context.

Consideration may need to be given to the following in an adaptive management context:

the need to capture rapid changes in temperature associated with rapidly varying flow releases downstream of impoundments;

the need to capture changes to different aspects of the thermal regime and at different times of day; and

the positioning and number of sampling locations downstream of the impoundment to capture and assess impact magnitude and recovery distance (including whether greater deviation from the above standards is permissible within the zone of initial mixing).

The literature relating to the ecological effects of metal precipitation downstream of hypolimnial releases is relatively anecdotal. It is possible that adverse effects could arise as a consequence of precipitated material (including manganese, iron, aluminium and in rare cases other metals, depending on the composition of reservoir sediments). However, this is considered likely to be restricted to the area immediately below the dam in the majority of cases. Mitigating summer cooling via implementation of surface or multi-level releases should also prevent release of anoxic water with high metal concentrations to the downstream watercourse. Nevertheless, consideration should be given to the following in an adaptive management context:

the need to collect samples from the bed sediment as well as from the water column; and

the positioning and number of sampling locations downstream of the impoundment to capture and assess impact magnitude and recovery distance.

It should be borne in mind that there may be considerable year-to year variation in

downstream effects.


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6. Knowledge gaps

The majority of literature relating to water quality (temperature and chemistry) effects of impoundments is directed towards the effects of hypolimnetic releases from large dams at high altitude; there is limited available literature relating to the adverse water quality impacts of lowland reservoirs, epilimnetic releases and/or small dams, as well as the downstream persistence of such effects.

The majority of literature relating to water quality (temperature and chemistry) effects on ecology downstream of impoundments relates to benthic macro-invertebrates and fish; there is limited available literature relating to the effects of water quality impacts downstream of impoundments on macrophytes, riparian birds, mammals, protected species (e.g. freshwater pearl mussels) or phytobenthos.


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7. Conclusions

The majority of the literature sources for this review were from North America and Europe; about a quarter were from the UK. This largely reflects the number impoundments in these regions, as well as their average size and scale of environmental impact. Few studies reported on the effects of impoundments on downstream water temperature and chemistry in isolation from the effects of river flows. Many studies suggest that riverine organisms respond to the interacting effects of water flow, temperature, chemistry and other physical parameters that are simultaneously affected by impoundments.

Large, deep impoundments (>5-10m maximum depth) with water released from near the bottom have the greatest effect on water temperature and chemistry in downstream water courses, with direct links to ecological responses. Shallower impoundments and those with surface water releases have less influence on the water quality and ecology of downstream water courses, except if they are anthropogenically eutrophic. This review has identified nine ecological risk factors associated directly with the effects of large impoundments on water temperature and chemistry in downstream water courses. Six risk factors relate to water temperature and three relate to water chemistry; all except one risk factor are associated with deep water releases:





Increase in mean and /or maximum summer temperatures (summer elevation) (small and/or shallow impoundments)




High concentrations of algae and chlorophyll a, with low concentrations of dissolved nutrients (that have been utilized by the algae) and high pH.

In the UK, summer cooling of rivers is considered to be the most important water temperature effect for macro-invertebrates and fishes downstream of impoundments. Summer cooling will slow down juvenile growth rates and disrupt the critical timing of insect emergence and oviposition. Such effects may be of more significance for coarse fish populations downstream of lowland reservoirs than for salmonid populations downstream of upland reservoirs, as coarse fish require warmer water temperatures for growth than salmonids. These effects might reduce the size of populations, but there is no evidence of populations being eliminated by water temperature effects downstream of impoundments in the UK, so these effects alone are not likely to drive WFD classification lower than Moderate in any given water body.

Effects on water temperature and chemistry and associated ecological impacts from large, deep reservoirs in the UK are likely to be relatively short-lived. In the UK, the majority of studies indicate a return to pre-impoundment conditions within ~20 km. In most situations, impacts are expected to be restricted to the first (and potentially second) WFD water bodies downstream of the impoundment.


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Some ecologically damaging thermal and chemical effects may also arise in rivers as a consequence of water releases from small/shallow impoundments in eutrophic or polluted reservoirs. The effects of releases from eutrophic reservoirs may persist for longer distances downstream, but the literature relating to such impacts is relatively sparse.

Overall, water temperature and chemistry effects of large impoundments have been reported to impact ecology in warm and continental climates. These effects have not been reported widely in cool temperate climates such as the UK, and where these effects occur, they are relatively localised to impoundment structures. The discharge regime of water from impoundments remains the primary driver for ecological effects downstream of most impoundments and most often will interact with any water quality effects. Water temperature and chemistry is likely to be the main driver for ecological effects downstream of UK impoundments only if they are very deep with hypolimnetic releases and/or eutrophic.

The risk assessment framework should be used to identify sites where water temperature and chemistry effects might have unintended negative consequences to ecology downstream of large impoundments when a flow release scheme is being considered as a mitigation measure for a HMWB. This framework sets out the process and questions that need to be asked using information and expert knowledge at a given site to assess the risk of impacts due to poor water quality and the likely effect this might have on WFD status.

The findings of the literature review indicate that existing WFD environmental standards are likely to remain appropriate in the context of cold water releases from impoundments. The literature suggests that a 2-3oC cooling effect during the summer months may have significant effects on ecology (particularly macro-invertebrates) and such changes are already broadly covered by existing standards.


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8. Recommendations

Where detailed information regarding stratification and downstream ecological impacts are available for a particular site, it is recommended that the UK regulators use such information to operate the decision framework presented in this report, to arrive at a site-specific solution. An adaptive management approach is recommended following implementation of any mitigation measures.

Consideration may need to be given to the following in an adaptive management context:

the need to capture rapid changes in temperature associated with rapidly varying flow releases downstream of impoundments;

the need to collect samples from the bed sediment as well as from the water column to assess impacts of metal deposition; and

the positioning and number of sampling locations downstream of the impoundment to capture and assess impact magnitude and recovery distance.

site-specific investigation of the potential impacts on use of changing the release depth on stratification processes is advisable for large, deep impoundments with an existing hypolimnial release (particularly where a change to a surface release is being considered).

It is important to recognise that the effects of changes in water chemistry and water temperature will act synergistically with changes in flow regime and sediment transport. At an ecosystem level it is unlikely to be possible to distinguish between the impacts of these effects, except in the most extreme cases. Indeed, in some cases it may be that addressing issues associated with flow and/or sediment transport may be of primary importance for driving downstream ecological improvements. It is therefore recommended that downstream water temperature and chemistry effects are considered in conjunction with effects on the flow and sediment transport regimes, as well as other anthropogenic impacts such as downstream habitat modifications.


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